[Federal Register Volume 79, Number 175 (Wednesday, September 10, 2014)]
[Rules and Regulations]
[Pages 53851-54123]
From the Federal Register Online via the Government Publishing Office [www.gpo.gov]
[FR Doc No: 2014-20814]



[[Page 53851]]

Vol. 79

Wednesday,

No. 175

September 10, 2014

Part II





Department of Commerce





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National Oceanic and Atmospheric Administration





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50 CFR Part 223





Endangered and Threatened Wildlife and Plants: Final Listing 
Determinations on Proposal To List 66 Reef-Building Coral Species and 
To Reclassify Elkhorn and Staghorn Corals; Final Rule

Federal Register / Vol. 79 , No. 175 / Wednesday, September 10, 2014 
/ Rules and Regulations

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DEPARTMENT OF COMMERCE

National Oceanic and Atmospheric Administration

50 CFR Part 223

[Docket No. 0911231415-4826-04]
RIN 0648-XT12


Endangered and Threatened Wildlife and Plants: Final Listing 
Determinations on Proposal To List 66 Reef-Building Coral Species and 
To Reclassify Elkhorn and Staghorn Corals

AGENCY: National Marine Fisheries Service (NMFS), National Oceanic and 
Atmospheric Administration (NOAA), Commerce.

ACTION: Final rule.

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SUMMARY: We, the National Marine Fisheries Service (NMFS), are 
publishing this final rule to implement our final determination to list 
the following 20 species as threatened: five in the Caribbean 
(Dendrogyra cylindrus, Orbicella annularis, Orbicella faveolata, 
Orbicella franksi, and Mycetophyllia ferox); and 15 in the Indo-Pacific 
(Acropora globiceps, Acropora jacquelineae, Acropora lokani, Acropora 
pharaonis, Acropora retusa, Acropora rudis, Acropora speciosa, Acropora 
tenella, Anacropora spinosa, Euphyllia paradivisa, Isopora 
crateriformis, Montipora australiensis, Pavona diffluens, Porites 
napopora, and Seriatopora aculeata) under the Endangered Species Act 
(ESA) of 1973, as amended. The two species currently listed as 
threatened (Acropora cervicornis and Acropora palmata) in the Caribbean 
still warrant listing as threatened. We also determined that a total of 
43 proposed species do not warrant listing as endangered or threatened 
species, and three proposed species are not determinable under the ESA. 
We have reviewed the status of the species and efforts being made to 
protect the species, and public comments received on the proposed rule, 
and we have made our determinations based on the best scientific and 
commercial data available. We also solicit information that may be 
relevant to the designation of critical habitat for the 20 species 
newly listed under this final rule.

DATES: The effective date of this final rule is October 10, 2014. 
Responses to the request for information regarding a subsequent ESA 
section 4(d) Rule and critical habitat designation must be received by 
November 10, 2014.

ADDRESSES: Submit responses to the request for information regarding a 
subsequent ESA section 4(d) Rule and critical habitat designation to 
National Marine Fisheries Service, Pacific Islands Regional Office, 
NOAA Inouye Regional Center, 1845 Wasp Blvd., Building 176, Honolulu, 
HI 96818; or National Marine Fisheries Service, Southeast Regional 
Office, 263 13th Avenue South, Saint Petersburg, FL 33701.

FOR FURTHER INFORMATION CONTACT: Lance Smith, NMFS, Pacific Island 
Regional Office, 808-725-5131; Jennifer Moore, NMFS, Southeast Regional 
Office, 727-824-5312; or Marta Nammack, NMFS, Office of Protected 
Resources, 301-427-8469. A list of the literature cited in this rule is 
available at http://coral.sero.nmfs.noaa.gov and http://www.fpir.noaa.gov/PRD/prd_coral.html.

SUPPLEMENTARY INFORMATION:

Background

    On October 20, 2009, the Center for Biological Diversity (CBD) 
petitioned us to list 83 reef-building corals as threatened or 
endangered under the Endangered Species Act (ESA) and designate 
critical habitat. The 83 species included in the petition were: 
Acanthastrea brevis, Acanthastrea hemprichii, Acanthastrea 
ishigakiensis, Acanthastrea regularis, Acropora aculeus, Acropora 
acuminata, Acropora aspera, Acropora dendrum, Acropora donei, Acropora 
globiceps, Acropora horrida, Acropora jacquelineae, Acropora listeri, 
Acropora lokani, Acropora microclados, Acropora palmerae, Acropora 
paniculata, Acropora pharaonis, Acropora polystoma, Acropora retusa, 
Acropora rudis, Acropora speciosa, Acropora striata, Acropora tenella, 
Acropora vaughani, Acropora verweyi, Agaricia lamarcki, Alveopora 
allingi, Alveopora fenestrata, Alveopora verrilliana, Anacropora 
puertogalerae, Anacropora spinosa, Astreopora cucullata, Barabattoia 
laddi, Caulastrea echinulata, Cyphastrea agassizi, Cyphastrea ocellina, 
Dendrogyra cylindrus, Dichocoenia stokesii, Euphyllia cristata, 
Euphyllia paraancora, Euphyllia paradivisa, Galaxea astreata, Heliopora 
coerulea, Isopora crateriformis, Isopora cuneata, Leptoseris 
incrustans, Leptoseris yabei, Millepora foveolata, Millepora tuberosa, 
Montastraea annularis, Montastraea faveolata, Montastraea franksi, 
Montipora angulata, Montipora australiensis, Montipora calcarea, 
Montipora caliculata, Montipora dilatata, Montipora flabellata, 
Montipora lobulata, Montipora patula, Mycetophyllia ferox, Oculina 
varicosa, Pachyseris rugosa, Pavona bipartita, Pavona cactus, Pavona 
decussata, Pavona diffluens, Pavona venosa, Pectinia alcicornis, 
Physogyra lichtensteini, Pocillopora danae, Pocillopora elegans, 
Porites horizontalata, Porites napopora, Porites nigrescens, Porites 
pukoensis, Psammocora stellata, Seriatopora aculeata, Turbinaria 
mesenterina, Turbinaria peltata, Turbinaria reniformis, and Turbinaria 
stellulata. Eight of the petitioned species occur in the Caribbean, and 
75 of the petitioned species occur in the Indo-Pacific region. Most of 
the 83 species can be found in the United States, its territories 
(Puerto Rico, U.S. Virgin Islands, Navassa, Northern Mariana Islands, 
Guam, American Samoa, Pacific Remote Island Areas), or its freely 
associated states (Republic of the Marshall Islands, Federated States 
of Micronesia, and Republic of Palau), though many occur more 
frequently in other countries.
    On February 10, 2010, we published a 90-day finding (75 FR 6616) 
that CBD had presented substantial information indicating the 
petitioned actions may be warranted for all of the petitioned species 
except for the Caribbean species Oculina varicosa. We also announced 
the initiation of a formal status review of the remaining 82 petitioned 
species, and we solicited input from the public on six categories of 
information: (1) Historical and current distribution and abundance of 
these species throughout their ranges (U.S. and foreign waters); (2) 
historical and current condition of these species and their habitat; 
(3) population density and trends; (4) the effects of climate change on 
the distribution and condition of these coral species and other 
organisms in coral reef ecosystems over the short and long term; (5) 
the effects of all other threats including dredging, coastal 
development, coastal point source pollution, agricultural and land use 
practices, disease, predation, reef fishing, aquarium trade, physical 
damage from boats and anchors, marine debris, and aquatic invasive 
species on the distribution and abundance of these coral species over 
the short- and long-term; and (6) management programs for conservation 
of these species, including mitigation measures related to any of the 
threats listed under No. 5 above.
    The ESA requires us to make determinations on whether species are 
threatened or endangered ``solely on the basis of the best scientific 
and commercial data available * * * after conducting a review of the 
status of the species * * * '' (16 U.S.C. 1533). Further, our 
implementing regulations

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specifically direct us not to take possible economic or other impacts 
of listing species into consideration (50 CFR 424.11(b)). We convened a 
Coral Biological Review Team (BRT) composed of seven Federal scientists 
from NMFS' Pacific Islands, Northwest, and Southeast Fisheries Science 
Centers, as well as the U.S. Geological Survey and National Park 
Service. The members of the BRT are a diverse group of scientists with 
expertise in coral biology, coral ecology, coral taxonomy, physical 
oceanography, global climate change, coral population dynamics and 
endangered species extinction risk evaluations. The BRT's 
comprehensive, peer-reviewed Status Review Report (SRR; Brainard et 
al., 2011) incorporates and summarizes the best available scientific 
and commercial information as of August 2011 on the following topics: 
(1) Long-term trends in abundance throughout each species' range; (2) 
potential factors for any decline of each species throughout its range 
(human population, ocean warming, ocean acidification, overharvesting, 
natural predation, disease, habitat loss, etc.); (3) historical and 
current range, distribution, and habitat use of each species; (4) 
historical and current estimates of population size and available 
habitat; and (5) knowledge of various life history parameters (size/age 
at maturity, fecundity, length of larval stage, larval dispersal 
dynamics, etc.). The SRR evaluates the status of each species, 
identifies threats to the species, and estimates the risk of extinction 
for each of the species out to the year 2100. The BRT also considered 
the petition, comments we received as a result of the 90-day finding 
(75 FR 6616; February 10, 2010), and the results of the peer review of 
the draft SRR, and incorporated relevant information from these sources 
into the final SRR. Additionally, we developed a supplementary, peer-
reviewed Draft Management Report (NMFS, 2012a) to identify information 
relevant to ESA factor 4(a)(1)(D), inadequacy of existing regulatory 
mechanisms, and protective efforts that may provide protection to the 
corals pursuant to ESA section 4(b).
    The response to the petition to list 83 coral species is one of the 
broadest and most complex listing reviews we have ever undertaken. 
Given the petition's scale and the precedential nature of the issues, 
we determined that our decision-making process would be strengthened if 
we took additional time to allow the public, non-Federal experts, non-
governmental organizations, state and territorial governments, and 
academics to review and provide information related to the SRR and the 
Draft Management Report prior to issuing our 12-month finding. Thus on 
April 17, 2012, we published a Federal Register notice announcing the 
availability of the SRR and the Draft Management Report, and 
specifically requested information on the following: (1) Relevant 
scientific information collected or produced since the completion of 
the SRR or any relevant scientific information not included in the SRR; 
and (2) relevant management information not included in the Draft 
Management Report, such as descriptions of regulatory mechanisms for 
greenhouse gas (GHG) emissions globally, and for local threats in the 
83 foreign countries and the United States, its territories (Puerto 
Rico, U.S. Virgin Islands, Navassa, Northern Mariana Islands, Guam, 
American Samoa, Pacific Remote Island Areas), or its freely associated 
states (Republic of the Marshall Islands, Federated States of 
Micronesia, and Republic of Palau), where the 82 petitioned coral 
species collectively occur. Further, in June 2012, we held listening 
sessions and scientific workshops in the Southeast region and Pacific 
Islands region to engage the scientific community and the public in-
person. During this public engagement period, which ended on July 31, 
2012, we received over 42,000 letters and emails. Also, we were 
provided with or we identified approximately 400 relevant scientific 
articles, reports, or presentations that were produced since the SRR 
was finalized, or not originally included in the SRR. We compiled and 
synthesized all relevant information that we identified or received 
into the Supplemental Information Report (SIR; NMFS, 2012c). 
Additionally, we incorporated all relevant management and conservation 
information into the Final Management Report (NMFS, 2012b). Therefore, 
the 82 candidate coral species comprehensive status review consists of 
the SRR (Brainard et al., 2011), the SIR (NMFS, 2012c), and the Final 
Management Report (NMFS, 2012b).
    On December 7, 2012, we published a proposed rule (77 FR 73219) to 
list 12 of the petitioned coral species as endangered (five Caribbean 
and seven Indo-Pacific) and 54 coral species as threatened (two 
Caribbean and 52 Indo-Pacific), and we determined 16 coral species (all 
Indo-Pacific) did not warrant listing as threatened or endangered under 
the ESA. This was the final agency action for those species which we 
determined were not warranted for listing. We also determined that two 
currently listed Caribbean corals (Acropora cervicornis and Acropora 
palmata) warranted reclassification from threatened to endangered. The 
findings in the proposed rule were based on the information contained 
within the reports described above (SRR, SIR, and Final Management 
Report). During a 90-day comment period, we solicited comments from the 
public, other concerned governmental agencies, the scientific 
community, industry, foreign nations in which the species occur, and 
any other interested parties on our proposal. We later extended the 
public comment period by 30 days, making the full comment period 120 
days. We received approximately 32,000 comments through electronic 
submissions, letters, and oral testimony from public hearings held in 
Dania Beach, FL; Key Largo, FL; Key West, FL; Rio Piedras, Puerto Rico; 
Mayaguez, Puerto Rico; Christiansted, St. Croix, U.S. Virgin Islands; 
Charlotte Amalie, St. Thomas, U.S. Virgin Islands; Hilo, Hawaii, HI; 
Kailua Kona, Hawaii, HI; Kaunakakai, Molokai, HI; Wailuku, Maui, HI; 
Lihue, Kauai, HI; Honolulu, Oahu, HI; Hagatna, Guam; Saipan, 
Commonwealth of the Northern Marianas Islands (CNMI); Tinian, CNMI; 
Rota, CNMI; Tutuila, American Samoa; and Washington, DC.
    During the public comment period, we received numerous comments on 
the proposed listing and the sufficiency or accuracy of the available 
data used to support the proposed listing determinations. In 
particular, comments raised questions and provided varied, often 
conflicting, information regarding the following topics:
    (1) The proposed species' listing statuses (e.g., certain species 
proposed as endangered should be threatened);
    (2) the sufficiency and quality, or lack thereof, of the species-
specific information used for each species' proposed listing 
determination;
    (3) the accuracy of the methods used to analyze the available 
information to assess extinction risk (including NMFS' ``Determination 
Tool'') and derive listing statuses for each of the proposed species;
    (4) the ability of corals to adapt or acclimatize to ocean warming 
and acidification;
    (5) the reliability, certainty, scale, and variability of future 
modeling and predictions of climate change; and
    (6) the effect local management efforts have on coral resilience.
    After considering these comments, we found that substantial 
disagreement existed regarding the sufficiency and accuracy of the 
available data used in support of the proposed determinations.

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As a result, we determined it was necessary to solicit additional data 
from those scientists who were identified by public comments and others 
who may have additional data to assist in resolving the substantial 
disagreement. Therefore, pursuant to the ESA section 4(b)(6)(B)(i), we 
determined that a 6-month extension of the deadline for final 
determinations on the proposed rule was necessary (78 FR 57835; 
September 20, 2013). We completed our data collection effort in the 
fall of 2013, and the relevant information that we received or 
collected was considered in the formulation of this final rule. The 
data collection effort was the final step in our thorough process to 
assemble the best available information on the status of the species 
addressed in this final rule. As a result, this final rule represents a 
logical evolution from the proposed rule, including some changes in our 
overall decision-making framework and a holistic reconsideration of the 
key elements that contribute to a species' listing status, as described 
in detail throughout this rule. Consequently, most of the listing 
determinations have changed between the proposed and final rules.

Listing Species Under the Endangered Species Act

    We are responsible for determining whether the 66 proposed coral 
species should be listed as threatened or endangered under the ESA, and 
whether the two species proposed for reclassification should be listed 
as endangered under the ESA (16 U.S.C. 1531 et seq.). Clonal, colonial 
organisms, such as corals, are vastly different in their biology and 
ecology than vertebrates, which are typically the focus of ESA status 
reviews. Therefore, concepts and terms that are typically applied to 
vertebrates have very distinct meanings when applied to corals. A 
`rare' coral may have millions of colonies as compared to a `rare' 
vertebrate, which may only have hundreds of individuals. To be 
considered for listing under the ESA, a group of organisms must 
constitute a ``species,'' which is defined in section 3 of the ESA to 
include ``any subspecies of fish or wildlife or plants, and any 
distinct population segment of any species of vertebrate fish or 
wildlife which interbreeds when mature.'' In the case of reef-building 
corals, the decision that a species is a listable entity is often 
complicated by several aspects of their biology including individual 
delineation, taxonomic uncertainty, identification uncertainty, and 
life history (e.g., colonialism and clonality).
    Section 3 of the ESA further defines an endangered species as ``any 
species which is in danger of extinction throughout all or a 
significant portion of its range'' and a threatened species as one 
``which is likely to become an endangered species within the 
foreseeable future throughout all or a significant portion of its 
range.'' Section 4(a)(1) of the ESA requires us to determine whether 
any species is endangered or threatened due to any one or a combination 
of the following five factors: (A) The present or threatened 
destruction, modification, or curtailment of its habitat or range; (B) 
overutilization for commercial, recreational, scientific, or 
educational purposes; (C) disease or predation; (D) the inadequacy of 
existing regulatory mechanisms; or (E) other natural or manmade factors 
affecting its continued existence. We are required to make listing 
determinations based solely on the best scientific and commercial data 
available after conducting a review of the status of the species and 
after taking into account efforts being made by any state or foreign 
nation to protect the species.
    This finding begins with an overview of coral biology, ecology, and 
taxonomy in the Corals and Coral Reefs section below, including whether 
each proposed species meets the definition of a ``species'' for 
purposes of the ESA. Specifically, are the proposed species 
determinable under the ESA given any discrepancies between their 
current morphologically-based taxonomy and any new genetic information 
that may result in taxonomic reclassification. Other relevant 
background information in this section includes the general 
characteristics of the habitats and environments in which the proposed 
species are found. The finding then summarizes information on factors 
adversely affecting and posing extinction risk to corals in general in 
the Threats Evaluation section. The Risk Analyses section then 
describes the framework applied to each of the species that resulted in 
final listing statuses for the proposed species. The Species-specific 
Information and Determinations section provides the best available 
species-specific information, which, coupled with the general portions 
of this final rule, provide the basis for the individual determinations 
for final listing status. Finally, we assessed efforts being made to 
protect the species and determined if these efforts are adequate to 
mitigate impacts and threats to the extent that a species does not meet 
one of the statutory statuses.
    Given the precedential and complex nature of this rule-making 
process, we took extra steps to assemble the best available information 
for informing the final listing determinations. Efforts to acquire this 
information first included the formation of an expert scientific panel 
(BRT) that used the best available scientific information at that time 
in a structured decision-making process to inform and write the SRR. 
Further, this process provided numerous opportunities for public input, 
including a public comment period after the 90-day finding in 2010 (75 
FR 6616; February 10, 2012), a unique public information-gathering 
period (77 FR 22749; April 17, 2012) prior to the release of the 
proposed rule in 2012, and a 120-day formal public comment period after 
the publication of the proposed rule. Finally, in a targeted data-
solicitation effort to resolve substantial scientific disagreement in 
the public comments on the proposed rule, we published a 6-month 
extension in September 2013 to gather additional information to further 
inform our final decisions (78 FR 57835; September 20, 2013). Over the 
course of this multi-year process, we gathered and reviewed thousands 
of scientific papers, journal articles, reports, and presentations 
(bibliography and select documents available at http://www.nmfs.noaa.gov/pr/species/invertebrates/corals.htm). In addition, we 
held a total of 19 public hearings in 2012 and 2013 throughout the 
Southeast and Pacific Islands regions, and received and reviewed over 
75,000 public comments during the information-gathering period in 2012 
and the proposed rule public comment period in 2012-2013, combined. 
These efforts ensure that this final rule is based upon the best 
available information on the proposed species at this time, as 
explained in more detail below.

Summary of Comments Received

    Below we address the comments received pertaining to the proposed 
listings or reclassifications of the 68 coral species in the December 
7, 2012, proposed rule (77 FR 73219). During the 120-day public comment 
period from December 7, 2012, to April 6, 2013, we received 1,120 
written and verbal responses (including public testimony during the 19 
public hearings). This included 1,119 unique comments on the proposed 
listings or reclassifications and 32,000 action alert responses in 
support of the rule organized by the petitioner CBD, which 
substantively constitutes one unique comment, and. The public comments 
received covered a wide breadth of topics, many of which

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were significant and within the scope of this rule-making. We 
summarized the comments, and these summaries and our responses are 
organized according to the sections of the proposed rule on which those 
comments were based. We have considered all public comments, and we 
provide responses to all relevant issues raised by comments. We have 
not responded to comments outside the scope of this rulemaking, such as 
comments regarding the potential economic impacts of ESA listings, 
comments suggesting that certain types of activities be covered in any 
future regulations pursuant to ESA section 4(d) for threatened species, 
or whether ESA listings are appropriate for species threatened by 
climate change. As explained in the Background above, this final rule 
was extended by 6 months to resolve substantial scientific disagreement 
in the public comments on six topics related to the proposed listing.

Comments on Taxonomic Uncertainty in Reef-Building Corals

    Comment 1: Many public comments on the proposed listing rule stated 
that species identification uncertainties and taxonomic uncertainties 
associated with many reef-building corals are problematic for the ESA 
listing determination process. Four comments specifically stated that 
the ability to determine the status of coral species under the ESA is 
impeded by the taxonomic uncertainty of many coral species. Two 
comments stated that genetic and genomic science is just beginning for 
corals, and as it develops it will likely show the current 
morphologically-based taxonomy is incorrect, completely changing 
current coral taxonomy. Therefore, management decisions based on the 
current taxonomy should be approached with caution. One comment stated 
that proper species identification, especially for the Indo-Pacific 
Acropora genus, is difficult and exacerbated by the use of outdated and 
inadequate information.
    Most of these comments are based on species identification 
uncertainties and the conflicting taxonomic results between recent 
genetics studies and traditional morphology-based taxonomy, and 
comments identified two potential problems: (1) Species identification 
and taxonomic uncertainty prevents many reef-building coral taxa, 
especially in the Indo-Pacific, from being determinable species under 
the ESA; and (2) even if these taxa are determinable species under the 
ESA, the taxonomic uncertainty confounds the available information 
regarding the status of each species, thus it is not possible to 
determine the listing status of these species with adequate confidence.
    Response: The comments correctly note that in some instances, lack 
of information, or ambiguity and uncertainty in available information, 
is so great that any listing determination on such a basis would be 
arbitrary. In our judgment, that is not the case for the proposed 
species, with a few exceptions noted below. The SRR concluded that the 
68 species in the proposed rule were determinable, including the 
species for which the SRR found that splitting or lumping petitioned 
species was necessary based on genetic studies. For the proposed rule, 
we agreed with the SRR, and considered the 68 species to be 
determinable for purposes of conducting a status review and determining 
listing status under the ESA.
    The public comments did not provide any studies or results, nor did 
we find any new studies or results, that significantly contradict the 
consideration of the traditional, morphologically described species as 
determinable species, with the exception of Pocillopora. We 
acknowledged in the proposed rule, however, that the taxonomic 
uncertainty for reef-building corals is not only real (Brainard et al., 
2011), but increasing in recent years as genetics studies have advanced 
(Stat et al., 2012; Veron, 2013). In the case of Pocillopora species, 
the taxonomic uncertainty has recently increased substantially such 
that the three proposed species in this genus are not determinable 
under the ESA (see Comment 2). For the remaining 65 species, the best 
available scientific information continues to support their 
classification as species. The taxonomic uncertainty associated with 
each species is considered along with other types of uncertainty when 
determining the status of each species in the Species-specific 
Information and Determinations section. In this way, the species 
identification and taxonomic uncertainty for each species is 
acknowledged and incorporated into each of the 65 determinations in 
this final rule.
    In this final rule, even though Millepora foveolata and Montipora 
lobulata were affirmed to be valid species, and there are few if any 
taxonomic uncertainty issues, the two species are so difficult to 
identify in the field that there is very little reliable information 
available for either species (Fenner, 2014b). Thus, as described in the 
Species-specific Information and Determinations below for M. foveolata 
and M. lobulata, the species identification uncertainty is so high for 
these species that there is not sufficient evidence to support listing 
determinations of threatened or endangered for either species. This is 
explained in more detail in each species' individual determination.
    Comment 2: Related to Comment 1, one comment identified Pocillopora 
as a problematic taxon and provided a recent scientific paper 
describing new genetic evidence of taxonomic contradictions between 
genetic and morphologic results for Pocillopora species (Pinz[oacute]n 
et al., 2013).
    Response: Based on information summarized in the SRR, the proposed 
rule split P. elegans into Indo-Pacific and Eastern Pacific nominal 
species, and proposed P. elegans (Indo-Pacific), P. elegans (Eastern 
Pacific), and P. danae for listing (P. danae only occurs in the Indo-
Pacific). However, after considering new information on taxonomic 
uncertainty throughout the genus Pocillopora that has become available 
since the publication of the proposed rule, including the paper 
(Pinz[oacute]n et al., 2013) submitted by the commenter, we no longer 
consider the three Pocillopora species that were proposed for listing 
to be determinable under the ESA. A range-wide phylogeographic survey 
that included most currently recognized pocilloporid species found that 
reliance on colony morphology is broadly unreliable for species 
identification, and that several genetic groups have highly limited 
geographic distributions. The study concluded that ``a taxonomic 
revision informed foremost by genetic evidence is needed for the entire 
genus'' (Pinz[oacute]n et al., 2013). Similarly, a phylogeographic 
survey of several currently recognized pocilloporid species 
representing a range of atypical morphologies thought to be rare or 
endemic to remote locations throughout the Indo-Pacific found that: (1) 
The current taxonomy of Pocillopora based on colony morphology shows 
little correspondence with genetic groups; (2) colony morphology is far 
more variable than previously thought; and (3) there are numerous 
cryptic lineages (i.e., two or more distinct lineages that are 
classified as one due to morphological similarities). The study 
concluded that ``the genus Pocillopora is in need of taxonomic revision 
using a combination of genetic, microscopic characters, and 
reproductive data to accurately delineate species'' (Marti-Puig et al., 
2013). Likewise, a more limited study of several currently recognized 
pocilloporid species in Moorea found that genetic groups do not 
correspond to colony morphology, and exhibit a wide range of 
morphological variation

[[Page 53856]]

(Forsman et al., 2013). These studies demonstrate that colony 
morphology in pocilloporids is a poor indicator of taxonomic 
relationships, for the following reasons: (1) Morphologically similar 
colonies may not be the same species (i.e., colonies of different 
species appear similar because of similar environmental conditions or 
other reasons); and (2) morphologically different colonies may be the 
same species (i.e., colonies of the same species appear different 
because of different environmental conditions or other reasons).
    While the current literature supports the taxonomic division of 
pocilloporids geographically into Indo-Pacific and Eastern Pacific 
groups, it indicates a high level of taxonomic uncertainty for all 
Pocillopora species that are found in both areas, such as P. elegans. 
Within these two geographic areas, colonies that resemble P. elegans 
may be different species, including possibly still undescribed species. 
That is, colonies may merely resemble P. elegans because of similar 
environmental conditions or other reasons, but actually may be 
different species. And the opposite type of taxonomic uncertainty also 
appears to be common, as colonies that do not resemble P. elegans may 
actually be P. elegans. That is, colonies that are P. elegans appear 
different because of different environmental conditions or other 
reasons (Forsman et al., 2013; Marti-Puig et al., 2013; Pinz[oacute]n 
et al., 2013). The recently appreciated taxonomic uncertainty is in 
addition to the historical morphological taxonomic uncertainty within 
the genus Pocillopora and for P. elegans specifically (Veron, 2013; 
Veron, 2014). While P. danae does not occur in the Eastern Pacific, 
similar taxonomic uncertainty problems occur for this species. That is, 
this species also had historical morphological taxonomic uncertainty 
(Veron, 2013), which has recently been compounded by genetic taxonomic 
uncertainty, leading Veron (2014) to conclude that the species likely 
requires a taxonomic revision. A new taxonomic revision of Pocillopora 
was published, in which P. danae was found to be a synonym of P. 
verrucosa, resulting in the traditional P. danae being included within 
P. verrucosa (Schmidt-Roach et al., 2014). However, the overall 
taxonomic uncertainty within Pocillopora, including for P. elegans and 
P. danae, has not been resolved, and in fact continues to increase as 
more studies are conducted. Thus, at this time, Pocillopora species are 
not determinable under the ESA. Therefore, we are withdrawing our 
proposal to list P. elegans (Indo-Pacific) as threatened, P. elegans 
(Eastern Pacific) as endangered, and P. danae as threatened; these 
species are not considered further in this final rule.
    Comment 3: Several comments objected to our agreement with the 
SRR's (Brainard et al., 2011) lumping of Montipora dilitata, M. 
flabellata, and M. turgescens into a single species, as well as the 
lumping of M. patula and M. verrilli into a single species, based on 
the results of a single genetics study by Forsman et al. (2010).
    Response: The objections in the public comments to lumping 
Montipora dilitata/M. flabellata/M. turgescens and M. patula/M. 
verrilli did not provide any new or supplemental information, nor did 
we find any new or supplemental information, contradicting the key 
study used by the SRR to consider these species as a group. We must use 
the best available science on which to base our determinations, and 
there is no indication that Forsman et al. (2010) is in error. However, 
as discussed in the response to Comment 1, we acknowledge that coral 
taxonomy is a rapidly growing field and that is creates uncertainty in 
determining a species under the ESA. This taxonomic uncertainty is 
considered in the individual Species-specific Information and 
Determination for the Montipora.

Comments on Reproductive Life History of Reef-Building Corals

    Comment 4: There were only a few comments related to the 
reproductive life history of corals. One comment stated that coral reef 
connectivity data are sparse, and while the majority of published 
studies on coral larval dispersal report evidence of local seeding and 
replenishment of reefs, other models and studies report sporadic 
periods of longer distance dispersal and recruitment events. The 
commenter felt that the proposed rule did not adequately address coral 
population dynamics and connectivity in determining the status of the 
candidate coral species under the ESA. Another comment stated that 
there is almost no information on any of the species' trends or 
recruitment rates, and the limited information available is based on 
qualitative opinion, not quantitative data. The comment also pointed 
out that the proposed rule agreed that the term `recruit' could be 
difficult to apply in the case of corals, which reproduce both sexually 
and asexually, and that the number of recruits per spawner depends on 
the age or size at which an entity is defined as a recruit. These 
comments assert that there is insufficient information on productivity 
and connectivity on which to base listing decisions.
    Response: Coral reproduction and connectivity are addressed 
generally in the Reproductive Life History of Reef-building Corals 
section. As each proposed coral species has a different reproductive 
life history, we more comprehensively address each species' 
reproduction, connectivity, and recruitment (when that information was 
available) as they relate to each species' status under the ESA in the 
Species-specific Information and Determinations section. The public 
comments did not provide any studies or information on reproduction or 
connectivity for any species except for Acropora cervicornis (see 
Species-specific Information and Determinations section). Any 
supplemental information we found is included in Species-specific 
Information and Determinations section.

Comments on Distribution and Abundance of Reef Building Corals

    Comment 5: We received several comments regarding the distribution 
and abundance of reef-building corals, mainly regarding the lack of 
species-specific information for many species' geographic distributions 
and population abundances. There were only a few comments related to 
determining the distribution and abundance of reef building corals, 
specifically on extrapolating individual corals to overall population 
abundance and distribution, on which to base a listing decision. One 
comment stated that coral population size and structure across the 
world's oceans is nearly impossible to determine with any accuracy 
because we use crude substitutes for individual animals in determining 
population and range information within a species. For example, there 
is a significant difference between using colony population and range 
estimates versus using polyp population and range estimates, which are 
essentially impossible to estimate. Another comment stated that it is 
not accurate to equate percent coral cover on reefs to population 
abundance (i.e., numbers of individuals). Any loss of coral cover often 
is manifest by loss of coral tissue over large portions of still living 
colonies, without the loss of the individual. Furthermore, it is 
unclear whether the loss of many separate but genetically-identical 
colonies (`clones') equates to the loss of a single but genetically-
distinct individual if some of the clone colonies survive. Another 
commenter noted that the distributions of the Indo-Pacific species are 
largely unknown due to their incredibly vast ranges encompassing 
numerous

[[Page 53857]]

archipelagos that include thousands of islands and atolls. The 
commenter emphasized this point by noting that there are between 30,000 
and 40,000 islands in Oceania which could potentially have populations 
of the proposed coral species. The comments described above 
collectively assert that listing decisions cannot be made due to the 
lack of species-specific information.
    Response: We acknowledge that it is difficult to quantify and 
qualify distribution and abundance for individual coral species. The 
ambiguity associated with the delineation of the individual in reef-
building corals is addressed in the Individual Delineation sub-section 
in the Corals and Coral Reefs section, including how we characterize 
the delineation of the individual for the species covered by this final 
rule. In response to public comments, we more adequately address each 
species' distribution and abundance as those characteristics relate to 
each species' determination status under the ESA in the Species-
specific Information and Determinations section. The public comments 
provided some useful information on the distribution and abundance of 
specific coral species, and we also collected supplemental information 
on distribution and abundance that is included in the Species-specific 
Information and Determinations section.

Comments on Coral Reefs, Other Coral Habitats, and Overview of 
Candidate Coral Environments

    Comment 6: Some comments asserted that the proposed rule focused 
too much on coral reefs rather than focusing on coral species. A couple 
of comments stated that corals thrive in places that are not coral 
reefs, even when nearby coral reefs are not thriving, underscoring the 
notion that reefs are not species. Another couple of comments stated 
that the focus on coral reefs and reef ecosystems, and the importance 
they have to reef-associated species, is improper for ESA listing 
analysis and added that NMFS cannot simply decide to treat reefs as a 
species under the ESA simply because evaluating reefs is easier.
    Response: The proposed rule acknowledged that reef-building coral 
species are not reef-dependent and provided a description of non-reefal 
habitats. Public comments did not provide information on how to 
interpret non-reefal habitat in our analysis, but in the Coral Habitats 
sub-section of this final rule we clarify the relevance of non-reefal 
habitats in determining each species' status under the ESA (e.g., 
providing variability in environmental conditions).
    Further, in the Coral and Coral Reefs section (Individual 
Delineation and Species Identification sub-sections), we explain that 
we define a coral species as the ``physiological colony'' (i.e., unit 
of the species that can be identified as an individual in the field) to 
ensure that we are evaluating the individual species and not coral 
reefs generally for determining ESA status. Public comments did not 
offer any information on how to define a coral species, but our 
explanations in the Individual Delineation and Species Identification 
sub-sections makes clear that we do not consider coral reefs as species 
in this final rule. However, it should be noted that defining an 
individual coral as the physiological colony in this final rule did not 
change how we interpreted abundance data for any species.
    Comment 7: A few comments stated that the proposed rule lacked 
species-specific information for mesophotic habitats (deep, lower-light 
areas, usually between 30 and 100 m deep). One comment stated that the 
coral communities of many Indo-Pacific jurisdictions have received 
little attention, with vast areas of reef remaining unexplored, 
especially for corals occurring in the mesophotic zone, which likely 
harbors populations of species that can also be found at shallower 
depths. Another comment stated that recent data from NOAA-supported 
studies of mesophotic reefs found these extensive and poorly studied 
ecosystems serve as refugia for numerous shallow water coral species, 
yet no survey data from these ongoing studies were included in the 
proposed rule. We also received two papers (Bridge and Guinotte, 2013; 
Kahng et al., 2014) that suggested the global diversity of some 
mesophotic corals may be underestimated and the biogeographic ranges of 
mesophotic corals are not fully explored.
    Response: The proposed rule briefly described mesophotic habitats 
and acknowledged that the amount of mesophotic habitat available is 
unknown and likely greater than the amount of shallow reef habitat. The 
proposed rule also stated there is greater coral cover on mesophotic 
reefs in the Indo-Pacific than in the Caribbean. However, more 
information has become available on this habitat type since publication 
of the proposed rule. Two papers (Bridge and Guinotte, 2013; Kahng et 
al., 2014) provided more information on the global diversity and 
biogeographic ranges of mesophotic corals and we have collected 
information on the magnitude and diversity of mesophotic habitat. The 
extent of mesophotic habitat is addressed in the Coral Habitats sub-
section. Mesophotic habitat's potential function as refugia for corals 
from ocean warming is addressed in the Spatial and Temporal Refugia 
sub-section. Where mesophotic habitat information is available for an 
individual coral species we have included and considered that 
information in the Species-specific Information and Determinations 
section.
    Comment 8: With regard to coral habitats being divided into only 
two global regions (i.e., Caribbean and Indo-Pacific), a couple of 
comments stated that the Indo-Pacific region was too coarse. 
Specifically, the comments stated that the Hawaiian Islands should be 
considered its own region or sub-region with Hawaiian species evaluated 
separately, due to Hawaii's isolated nature and significant number of 
endemic species.
    Response: We recognize that there may be numerous distinct sub-
regions throughout the Caribbean and Indo-Pacific basins for some or 
all species, and that some coral species are endemic to Hawaii. 
However, under the ESA, we must evaluate the status of the species 
throughout their entire ranges. Invertebrate species, such as corals, 
cannot be divided further into Distinct Populations Segments (DPS) 
under the ESA, since DPS specifically refer only to vertebrate species. 
Therefore, we cannot identify sub-regions, such as Hawaii, as its own 
distinct geographic range and evaluate the status of more broadly 
distributed species only within that specific area. In addition, as 
described in the Risk Analyses--Statutory Standard sub-section of this 
final rule, we were not able to identify a significant portion of its 
range (SPOIR) for any of the proposed corals and therefore could not 
evaluate whether the status of the species within that portion of its 
range impacts the overall status of the species throughout its range.
    Comment 9: We received a few comments regarding the consideration 
and inclusion of Traditional Ecological Knowledge (TEK), particularly 
from local island cultures (Hawaiian, Chamorro, and Samoan), as best 
available information for our listing determination process. One 
comment noted the importance of corals and coral reefs to island 
cultures in the Pacific Islands region, in particular to native 
Hawaiians. The comment criticized the lack of TEK in the SRR and 
proposed rule for the candidate corals, stating that coral biology and 
ecology is a fundamental part of TEK, and that their TEK is part of 
best available science.

[[Page 53858]]

    Response: We agree that TEK provides an important and unique 
perspective on local ecosystems, their status, threats, and changes 
over time; when relevant information was made available to us, we 
incorporated it into the proposed rule. We also acknowledge that this 
information is not necessarily accessible in academic peer reviewed 
journals or text books. Therefore, we requested any additional TEK-
related information on the biology, ecology, threats, and extinction 
risks of the 65 coral species on numerous occasions for inclusion 
within this final rule. While we received public comments and listened 
to several public testimonies from community members in both the 
Pacific Islands and Southeast regions that disagreed with our proposed 
listing determinations, we did not receive any TEK-related information 
or data on the biology, ecology, threats, or extinction risks for any 
of the 65 coral species within this final rule.

Comments on Threats Evaluation

    Comment 10: We received a large number of public comments on the 
various threats to corals and coral reefs. In addition to the specific 
comments on the nine most important threats, one comment stated that 
there should be no doubt that corals and coral reefs throughout the 
world are in serious trouble and in decline due to the effects of 
anthropogenic stressors. Another commenter asked whether the mere 
threats from anthropogenic impacts are sufficient for ESA listing. Yet 
another commenter requested that recreational boating activities should 
be recognized as a specific threat, even though recreational boating 
activities may only present a relatively minor risk to coral species.
    Response: As described in the proposed rule, there are nine threats 
considered to be the most significant to the current or expected future 
extinction risk of reef-building corals. The comments and responses on 
these nine threats (ocean warming, disease, ocean acidification, 
trophic effects of fishing, sedimentation, nutrients, sea-level rise, 
predation, and collection and trade) are addressed individually below. 
We acknowledged that recreational boating activities may present some 
risk to coral species and it was included in the description of the 
threat ``Human-induced Physical Damage'' in the SRR. However, we 
determined that threat's contribution to the extinction risk of corals, 
generally, is negligible to low.
    We also recognized that anthropogenic threats are affecting coral 
species worldwide and may be sufficient for an ESA listing if the 
species meets the definition of threatened or endangered. That is, if 
the species is currently in danger of extinction or may become so in 
the foreseeable future due to any one or a combination of the five 
factors under Section 4 of the ESA (in which the various threats are 
categorized) then the species may be listed.

Comments on Global Climate Change--General Overview

    Comment 11: We received many comments on the general treatment of 
global climate change in the proposed rule and supporting documents. 
The Global Climate Change--General Overview section in the proposed 
rule and the global climate change portion of the SRR describe past, 
current, and future GHG emissions and atmospheric concentrations and 
the associated past, current, and future general effects on coral reef 
ecosystems, based primarily on the International Panel on Climate 
Change's (IPCC) Fourth Assessment Report (AR4), The Physical Basis 
(IPCC, 2007) and supporting literature.
    Some comments stated that we did not adequately account for the 
uncertainty in climate change modeling. A few comments stated that 
global temperature has been stable for the last ten years or that 
warming has slowed down since 2000. One commenter provided two recent 
papers (Guemas et al., 2013; Hansen et al., 2012) that showed global 
mean surface temperatures did not increase as much as had been 
predicted from 2000 to 2010.
    Some comments stated that GHG emissions and global temperatures 
continue to rise unabated. One comment referenced two studies (Frieler 
et al., 2012; van Hooidonk et al., 2013b) that projected the frequency 
of coral reef bleaching under different levels of warming and emissions 
scenarios, indicating that significant and immediate GHG reductions are 
critical to prevent coral reefs from degradation and collapse. Another 
comment also referenced van Hooidonk et al. (2013b) and stated that 
targets for atmospheric carbon dioxide (CO2) concentrations 
must be lower than 450 parts per million (ppm) to protect coral reef 
ecosystems. Yet another comment stated that scientific modeling 
indicates that within 40 to 50 years, reef decline will pass a tipping 
point, largely due to the increasing impacts of climate change, and may 
not be reversible over ecological time scales. Another comment pointed 
out that climate change also could likely increase corals' exposure to 
cold water stress, which studies have shown can cause extensive 
mortality of corals (Colella et al., 2012; Schopmeyer et al., 2012).
    Response: We agree with commenters and acknowledge that there is 
uncertainty associated with climate change projections. Climate change 
projections over the foreseeable future are associated with three major 
sources of uncertainty: (1) The projected rate of increase for GHG 
concentrations; (2) strength of the climate's response to GHG 
concentrations; and (3) large natural variations. The recent warming 
slow-down is an example of a large natural variation that was not 
anticipated by previous models. However, AR4's projections were built 
upon scientifically accepted principles, which fairly simulated many 
large scale aspects of present-day conditions, providing the best 
available information on climate change at the time the proposed rule 
was published. The IPCC's Fifth Assessment Report (AR5), Climate Change 
2013: The Physical Science Basis (IPCC, 2013), commonly referred to as 
the Working Group I Report (WGI) became available in September 2013, 
and supersedes AR4; accordingly, this final rule relies on the 
information provided in AR5's WGI. Despite the advance of climate 
change science in recent years, there is still complexity and 
uncertainty associated with projections of global climate change. 
However, the current state of climate change science is capable of 
producing informative projections that provide a rational basis for 
considering likely patterns in future climate change-related threats to 
reef-building corals. More detail on the overall complexity associated 
with projections of global climate change, major sources of uncertainty 
in climate change projections, and a summary of AR5's WGI, including 
the pathway that we consider the most impactful to corals, are 
addressed in Threats Evaluation--Global Climate Change Overview sub-
section.
    We also acknowledge the observed recent hiatus/slow-down in the 
rate of global surface air temperature increase, and we have 
accordingly provided a description of the hiatus/slowdown and its 
implications in the Threats Evaluation--Ocean Warming sub-section. In 
summary, despite unprecedented levels of GHG emissions in recent years, 
a slow-down in global mean surface air temperature warming has occurred 
since 1998, which AR5's WGI refers to as a ``hiatus.'' Despite this 
slowdown in warming, the period since 1998 is the warmest recorded and 
``Each of the last three decades has been successively warmer at the 
Earth's surface than any preceding decade since

[[Page 53859]]

1850.'' The slow-down in global mean surface warming since 1998 is not 
fully explained by AR4 or AR5 WGI's models, but is consistent with the 
substantial decadal and interannual variability seen in the 
instrumental record and may result, in part, from the selection of 
beginning and end dates for such analyses.
    Public comments provided supplemental information on several 
aspects of global climate change, as described above. We also collected 
information to inform how we assess the effects of global climate 
change to corals, including the IPCC Working Group II report on 
impacts, adaptation, and vulnerability. We maintain that global climate 
change is central to assessing extinction risk for the corals in this 
final rule. As described in more detail in the Threats Evaluation--
Global Climate Change Overview sub-section below, the supplemental 
information underscores the complexity and uncertainty associated with 
projecting the extent and severity of effects of global climate change 
across the ranges of reef-building corals.

Comments on Ocean Warming (High Importance Threat, ESA Factor E)

    Comment 12: We received several comments on general future 
projections of ocean warming levels. One commenter stated that climate 
change models applied in our assessment are too coarse to accurately 
predict the conditions reefs will experience in the future and that 
real conditions are impacted by bathymetry, water mixing, wind 
patterns, fresh water inputs, and other bio-geographic factors. The 
commenter concluded that existing projections for sea surface 
temperature are not sufficient to conclude the species face an 
existential threat. Other comments also criticized the use of AR4's 
worst-case scenario as the basis for determining the most likely future 
scenario with regard to ocean warming, and related topics such as the 
proposed rule's lack of consideration for the post-1998 hiatus in 
global warming.
    Response: In the proposed rule, we discussed the numerous, complex 
spatial and temporal factors that compound uncertainty associated with 
projecting effects of ocean warming on corals in the future, and we 
have determined that ocean warming will not affect all species in all 
locations uniformly over the foreseeable future. We believe that 
different bio-geographic factors such as bathymetry, water mixing, wind 
patterns, and fresh water will likely impact conditions corals will 
experience over the foreseeable future. We also recognized that global 
climate change models are associated with uncertainty, as discussed in 
response to comment 11 above. However, in response to comments on ocean 
warming projections, such as criticism of the reliance of the proposed 
rule and supporting documents on AR4 (IPCC, 2007) and the lack of 
consideration of the ocean warming hiatus, we provide a review of the 
best available information on these topics, including AR5's WGI Report 
(IPCC, 2013), in the Threats Evaluation--Global Climate Change 
Overview, Representative Concentration Pathways (RCP) 8.5 Projections, 
and Ocean Warming sub-sections below. These data support the conclusion 
in the proposed rule that ocean warming is increasing in severity, and 
is likely to continue increasing in severity within the ranges of reef-
building corals. However, a key difference between the proposed and 
final rule is that we now more fully consider the ability of each 
species' spatial and demographic traits to moderate exposure to 
threats, including warming, and place appropriate emphasis on the non-
uniform nature of global threats at the regional and local levels that 
allows habitat heterogeneity to play a role in buffering a species 
against vulnerability to extinction. The significance of coral 
abundance and distribution, and habitat heterogeneity, to this final 
rule is described in more detail in the Corals and Coral Reefs, Risk 
Analyses and Species-specific Information and Determinations sections 
of this rule.
    After reviewing the public comments and information provided in 
AR5's WGI our conclusion regarding the threat of ocean warming remains 
unchanged from the proposed rule. We maintain that ocean warming is a 
high importance threat in assessing global extinction risk for the 
corals in this final rule, while we also acknowledge that the 
interpretation of future climate change threats to corals is associated 
with complexity and uncertainty, and that effects on individual species 
of reef-building corals are difficult to determine as described in more 
detail in the Threats Evaluation--Global Climate Change Overview 
subsection below.
    Comment 13: Many comments criticized the proposed rule for not 
accounting for spatial variability in ocean warming and overlooking 
regional and local variability in conditions leading to warming-induced 
bleaching, which may be more or less severe regionally or locally than 
the overall warming. For example, we received two comments requesting 
us to review the literature for information regarding current and 
projected regional differences in sea surface temperature anomalies and 
for variations in the responses of individual coral species across 
their ranges. Comments noted that coral species and their symbionts are 
not uniformly susceptible and/or resilient to climate change across 
their ranges. That variability results in heterogeneous responses of 
coral species to ocean warming both in different parts of the ranges 
and also at different rates in the future. Another comment provided 
information from van Hooidonk (2013b) regarding spatial and temporal 
variability of ocean warming within different regions. The commenter 
identified reef locations that appear to be less vulnerable to 
bleaching, including the southern Great Barrier Reef (GBR), the western 
Indian Ocean, Persian Gulf, Red Sea, Thailand, New Caledonia and French 
Polynesia, as well as other locations that appear to be more vulnerable 
to bleaching, including the western Pacific warm pool, northwestern 
Australia, west Papua New Guinea and the central Pacific islands of 
Tokelau. Another commenter stated that the corals at Flower Garden 
Banks National Marine Sanctuary seem to be less affected by elevated 
sea surface temperatures that are impacting corals in other parts of 
the wider Caribbean.
    Response: We discussed spatial (i.e., regional and/or local) 
variability of ocean warming impacts to corals in the proposed rule and 
we agree that ocean warming will not affect all species in all 
locations uniformly over the foreseeable future, and that different 
regions are predicted to experience the effects of ocean warming on 
different time scales and at different magnitudes than others. We 
provide a review of all the best available information on spatial 
variability in ocean warming, including any information provided via 
public comment or gathered ourselves since the proposed rule was 
published, in the Threats Evaluation--Global Climate Change Overview, 
RCP8.5 Projections, and Ocean Warming sub-sections below. These data 
support the conclusion in the proposed rule that ocean warming is 
increasing in severity, and likely to continue increasing in severity 
within the ranges of reef-building corals. This review also underscores 
the complexity and uncertainty associated with spatial variability in 
ocean warming across the ranges of reef-building corals. A key 
difference between the proposed and final rule is that we now more 
fully consider the ability of each species' spatial and demographic 
traits to moderate exposure to threats, including warming, and place 
appropriate emphasis on the non-uniform nature of

[[Page 53860]]

global threats at the regional and local levels which allows habitat 
heterogeneity to play a role in buffering a species against 
vulnerability to extinction. The significance of coral abundance and 
distribution and habitat heterogeneity to this final rule is described 
in more detail in the Corals and Coral Reefs, Risk Analyses and 
Species-specific Information and Determinations sections of this rule.
    Comment 14: Comments on the overview of ocean warming and coral 
reefs focused on projected effects of ocean warming on coral reef 
ecosystems, rather than on reef-building coral species. These comments 
comprise two distinct views. Some comments emphasized that coral reefs 
are likely to decline sharply in the future because of increasing GHG 
emissions, while other comments emphasized that recent reviews indicate 
a wide range of possible responses by coral species. For example, one 
commenter cited Frieler et al. (2012) and stated that the estimated 
frequency of coral bleaching at different levels of global warming 
showed that limiting warming to 1.5 [deg]C above pre-industrial levels 
is unlikely to protect most of the world's reefs from degradation. The 
commenter further explained that even under the lowest of the IPCC AR5 
emissions scenarios (RCP3-PD) and optimistic assumptions regarding 
thermal adaptation, approximately one-third (range from 9 to 60 
percent) of the world's coral reefs will experience long-term 
degradation. Another commenter cited Donner (2009) and similarly stated 
that the projected increase in sea surface temperatures due to the 
physical commitment from the present accumulation of GHGs due to 
anthropogenic activity, as well as the amount of GHGs likely to be 
emitted, is sufficient to cause frequent and higher magnitude heat 
stress for the majority of the world's coral reefs by 2050. Another 
commenter provided information from Kiessling et al. (2004) and 
Carpenter et al. (2008) and asserted that if bleaching events become 
very frequent, many species may be unable to maintain breeding 
populations as repeated bleaching causes potentially irreversible 
declines, perhaps mimicking conditions that led to previous coral 
extinctions. In contrast, some commenters disagreed with our conclusion 
of the projected effects of ocean warming on corals and coral reef 
ecosystems in the proposed rule. As described above in Comment 13, many 
commenters pointed out several studies showing regional and local 
variability in responses of corals and coral reefs to ocean warming.
    Response: We summarized the best available information on the 
interaction between ocean warming and corals reefs in the proposed 
rule, and concluded that ocean warming is a severe and increasing 
threat to corals. The public comments and supporting papers we received 
on the overview of ocean warming and coral reefs generally support the 
conclusion in the proposed rule that ocean warming is an important and 
increasing threat to coral reefs. However, the other comments 
underscore the uncertainty associated with projecting the effects of 
ocean warming on coral reefs in the future, and as described in our 
response to Comment 13, we also acknowledge that there is and will 
continue to be regional and local variability in responses of corals to 
ocean warming over the foreseeable future. We acknowledge that ocean 
warming will not act uniformly on all species at all times over the 
foreseeable future. Further, we recognize that the responses of each 
species to ocean warming will vary across their ranges over the 
foreseeable future. Additionally, as described in previous comment 
responses, a key difference between the proposed and final rule is that 
we now more fully consider the threat-buffering capacity of each 
species' unique characteristics, and place appropriate emphasis on the 
non-uniform nature of global threats at the regional and local levels 
which allows habitat heterogeneity to play a role in buffering a 
species against vulnerability to extinction.
    Comment 15: We received comments on specific effects of ocean 
warming on reef-building corals that covered various topics, including 
the interactions of warming-induced bleaching with other threats. For 
example, one commenter noted that anthropogenic climate change (e.g., 
ocean warming) weakens coral colonies and renders them more susceptible 
to disease, which is also covered in the Threats Evaluation--Disease 
sub-section below. Other commenters also emphasized the potential for 
ocean warming to act synergistically with other threats such as 
nutrification as well as overfishing. Another commenter provided 
information from Ferrier-Pag[egrave]s et al. (2010) suggesting 
remarkable tolerance to global change, such as the potential to reduce 
bleaching vulnerability through increased feeding rates.
    Response: In the proposed rule, we discussed how multiple threats 
stress corals simultaneously or sequentially, whether the effects are 
cumulative (the sum of individual stresses) or interactive (e.g., 
synergistic or antagonistic). The comments and supporting papers we 
received on these topics provide supplemental information (such as 
synergistic effects of ocean warming with other threats), which has 
been incorporated and considered in our assessment, as described in 
more detail in the Threats Evaluation--Ocean Warming sub-section. The 
comments and supporting papers support the conclusion in the proposed 
rule that the impacts of ocean warming on reef-building corals are 
increasing in severity and likely to continue increasing in severity. 
This information also underscores the great complexity and high 
uncertainty associated with the various specific effects of ocean 
warming, including synergistic effects with other threats, across the 
ranges of reef-building corals. We continue to acknowledge that 
susceptibility of a species to a threat depends on the combination of: 
(1) Direct effects of the threat on the species; and (2) the cumulative 
and interactive (synergistic or antagonistic) effects of the threat 
with the effects of other threats on the species. In the proposed rule, 
we considered how the cumulative or interactive effects altered the 
rating assigned to a threat susceptibility in isolation. However, upon 
further consideration, we need to evaluate the extent to which one 
threat influences the susceptibility of an individual species to 
another threat with more species-specific information, in connection 
with all the other elements that influence a species' extinction risk. 
Generally, cumulative and interactive processes are complex and 
uncertain and existing information about threats interactions is only 
based on a few studies on a few species. Where possible, when we have 
species-specific or applicable genus-level information on cumulative or 
interactive effects, we have applied this information to that 
particular species' susceptibilities in a more integrated manner.
    Comment 16: We received several comments on the capacity of reef-
building corals for acclimatization and adaptation to ocean warming, 
covering various specific characteristics of reef-building corals that 
may contribute to such capacity. Mostly, commenters asserted that we 
did not adequately consider the ability of corals to acclimatize or 
adapt to changing temperatures. Several comments cited empirical 
evidence that corals have already adapted to ocean warming, thereby 
demonstrating the potential for acclimatization or adaptation. For 
example, one comment letter provided information from Pandolfi et al. 
(2011) and Cahill et al. (2013) stating that more recent analyses 
incorporating thermal

[[Page 53861]]

tolerance of species indicate a wide range of outcomes including 
maintenance of comparable levels of cover to 2100 and beyond. Another 
commenter provided data from Maynard et al. (2008) and Guest et al. 
(2012) showing that many types of coral show surprisingly large (~0.5-1 
[deg]C) increases in thermal tolerance after a single mass bleaching 
event, due to either adaptation or acclimatization. In another comment 
letter, information provided from Jones and Berkelmans (2010) and Baker 
et al. (2004) show that the acclimatization potential of corals to 
increased temperatures is an active area of research, with a focus on 
identifying heat-resistant phenotypes. Another commenter pointed to the 
coral species that occur in the Arabian Gulf as an example of species 
adapting to warmer temperatures.
    Response: In the proposed rule we acknowledged that there is some 
evidence to suggest that reef-building corals may have various 
mechanisms for acclimatization and adaptation to ocean warming. These 
topics were described in the Ocean Warming sub-section of the proposed 
rule, and we concluded that existing scientific information was 
inconclusive on how these processes may affect individual corals' 
extinction risk, given the projected intensity and rate of ocean 
warming. The public comments and supporting papers have been 
incorporated and considered in our assessment, as described in more 
detail in the Threats Evaluation--Ocean Warming sub-section and the 
Species-specific Information and Determinations section. However, the 
supplemental information does not alter the conclusion in the proposed 
rule that the capacity for acclimatization and adaptation of reef-
building corals to ocean warming is inconclusive for corals generally 
at this time.

Comments on Disease (High Importance Threat, ESA Factor C)

    Comment 17: One comment regarding the decline of Caribbean coral 
populations cited land-use changes as well as disease outbreaks (among 
other local threats) as the causes of Caribbean coral decline rather 
than climate change. Some comments also provided such information 
pertaining to specific species. For example, one comment stated that 
the genetic diversity of Acropora cervicornis in Florida may be 
sufficient to maintain viability and resilience to environmental 
perturbations and disease.
    Response: The proposed rule described how disease had a major role 
in the initial decline of Caribbean coral populations as described in 
the Coral Reefs, Other Coral Habitats, and Overview of Candidate Coral 
Environments sections of the proposed rule. Further, in the Threats 
Evaluation--Disease section of this rule, we acknowledge diseases are 
of high importance with regard to extinction risk of corals. However, 
in assessing extinction risk over the foreseeable future, climate 
change-related threats are highly important to all reef-building 
corals. Any species-specific information provided on disease is 
included in the Species-specific Information and Determinations section 
later in this rule.
    Comment 18: One commenter noted the explicit link between coral 
bleaching, disease, and the larger driving environmental factor of 
climate change by citing several studies that show anthropogenic 
climate change weakens coral colonies and renders them more susceptible 
to disease (Harvell et al., 1999; Harvell et al., 2002; Knowlton, 
2001). Another commenter provided information from Muller and van 
Woesik (2012), stating that exceeding environmental disease thresholds 
will most likely become increasingly common in rapidly warming oceans, 
leading to more frequent coral-disease outbreaks. The study suggested 
that that the expression of some coral diseases occurs when (1) 
environmental thresholds are exceeded and (2) these environmental 
conditions either weaken the corals, which are then more susceptible to 
infection, or increase the virulence or abundance of pathogens. In 
other words, corals that experience bleaching are more likely to suffer 
from disease outbreaks and subsequent mortality.
    Response: In the proposed rule, we described the importance of 
disease as a threat to corals and the potential for disease to act 
synergistically with other threats such as ocean warming. We also 
understand that assessing the threat of disease is highly complex, as 
the cause or causes of many coral diseases remains either unknown or 
poorly understood. Overall, the public comments we received underscored 
and supported the analysis in the SRR and the proposed rule. In 
addition to public comments, we collected a significant amount of 
information on disease that became available since the proposed rule 
published. Thus, we maintain that disease is a high importance threat 
to the extinction risks of the 65 corals in this final rule. All of the 
supplemental information received or otherwise collected has been 
detailed and summarized in the Threats Evaluation--Disease sub-section 
of this final rule. The extent to which the extinction risk of a 
particular coral species is impacted by disease is discussed in more 
detail in the Species-specific Information and Determinations section 
below.

Comments on Ocean Acidification (Medium-High Importance Threat, ESA 
Factor E)

    Comment 19: We received public comments on the description of and 
future projections of ocean acidification, which provided information 
on the complexity of ocean chemistry on corals, and criticism of the 
use of the AR4's worst-case scenario as the basis for determining the 
most likely future scenario with regard to ocean acidification. For 
example, one commenter asserted that global projections of ocean 
acidification are too coarse and do not take into consideration 
competing and extremely localized factors that affect local 
CO2 concentrations (e.g., local atmospheric processes, local 
biological processes, local temperature, and upwelling from deeper 
waters). The commenter emphasized that despite acknowledging the 
multitude of local, regional, and seasonal factors that may cause local 
CO2 concentrations to increase and pH to decrease, we opted 
instead to base our reef-scale threat analysis on generalized 
acidification predictions from global models. Other commenters also 
criticized our reliance on the IPCC's AR4 report as the basis for our 
threat evaluation of ocean acidification to corals.
    Response: In the proposed rule we acknowledged that numerous, 
complex spatial and temporal factors compound uncertainty associated 
with projecting effects of ocean acidification on corals in the future. 
We also acknowledged that global climate change models are associated 
with uncertainty. We further acknowledge that the interpretation of 
future climate change threats to corals is complex and that effects on 
individual species of reef-building corals are difficult to determine, 
as described in more detail in the Threats Evaluation--Global Climate 
Change Overview subsection. However, we agree with commenters that 
ocean acidification will not affect all species in all locations 
uniformly over the foreseeable future, and that different locations 
will experience the effects of ocean acidification at different time 
scales and at different magnitudes than others. We provide a review of 
all the best available information, including a review of AR5's WGI 
(IPCC, 2013) in the Threats Evaluation--Global Climate Change Overview, 
RCP8.5 Projections, and Ocean Acidification sub-sections. Upon review 
of the information provided in AR5's WGI and public comments, our

[[Page 53862]]

conclusion regarding the threat of ocean acidification remains 
unchanged from the proposed rule. We maintain that ocean acidification 
is increasing in severity, and is likely to continue increasing in 
severity, within the ranges of reef-building corals, and is a medium-
high importance threat in assessing extinction risk for the 65 corals 
in this final rule. However, as described in earlier comment responses, 
a key difference between the proposed and final rule is that we now 
more fully consider the ability of each species' spatial and 
demographic traits to moderate the impacts of threats, and we place 
appropriate emphasis on the non-uniform nature of global threats at the 
regional and local levels which allows habitat heterogeneity to play a 
role in buffering a species against vulnerability to extinction.
    Comment 20: We received a comment regarding variability in ocean 
acidification on coral reefs related to fluctuations in pH from 
localized factors such as seagrass beds. The commenter provided 
information from Manzello et al. (2012) indicating that local and 
regional biochemical processes buffer effects of ocean acidification in 
locations such as the Gulf of Mexico and South Atlantic. Manzello et 
al. (2012) reported that the photosynthetic uptake and sequestering of 
carbon dioxide by seagrasses and other macroalgae and the positive 
growth response by seagrasses to increasing dissolved carbon dioxide 
(Palacios and Zimmerman, 2007) may create ocean acidification refugia 
for corals. Comments on specific effects of ocean acidification on 
coral reefs and reef-building corals focused on capacity for 
acclimatization of corals to acidification, and evidence that some 
coral species are resistant to low pH.
    Response: In the proposed rule, we discussed that numerous, complex 
spatial and temporal factors compound uncertainty associated with 
projecting effects of ocean acidification on corals and coral reefs in 
the future, and we agree with the comment that ocean acidification will 
not affect all species in all locations uniformly over the foreseeable 
future, and that different locations will experience the effects of 
ocean acidification at different time scales and at different 
magnitudes than others. In response to comments on spatial variability 
of ocean acidification, such as lack of consideration of localized 
increase in pH from adjacent seagrass beds, we provide a review of the 
best available information on spatial variability in ocean 
acidification, including any information provided by public comments as 
well as any information we gathered ourselves since the proposed rule 
was published, in the Threats Evaluation--RCP8.5 Projections and Ocean 
Acidification sub-sections. These data in our view still support the 
conclusion in the proposed rule that ocean acidification is increasing 
in severity, and likely to continue increasing in severity within the 
ranges of reef-building corals; however, as described in earlier 
comment responses, a key difference between the proposed and final rule 
is that we now more fully consider the threat moderation capacity of 
each species' spatial and demographic traits, and of habitat 
heterogeneity.
    Comment 21: We received one comment that identified a couple of 
ocean acidification and coral reef calcification rate studies that were 
not included in the SRR and proposed rule. The commenter provided two 
studies: One showing that coral calcification increases with global 
warming (McNeil et al., 2004), and another study showing that corals 
are already thriving in conditions similar to the ocean acidification 
conditions predicted by the IPCC for 2100 (Hofmann et al., 2011).
    Response: In the proposed rule and supporting documents we 
acknowledged that some exceptional areas exist where reef-building 
coral communities appear to be thriving under naturally high 
CO2 concentrations. As described in the comment response 
above to Comment 19, we agree that ocean acidification will not act 
uniformly on all species in all locations over the foreseeable future. 
We provide a review of all the best information available on the threat 
of ocean acidification, including these studies, which we received in 
public comments, and any information we gathered ourselves in the 
Threats Evaluation--Ocean Acidification sub-section (e.g., Shamberger 
et al., in press). This supplemental information supports the proposed 
rule's conclusion that the threat of ocean acidification has already 
impacted corals and coral reefs and will become increasingly severe 
from now to 2100, with increasingly severe consequences for corals and 
coral reefs. However, as described in previous comment responses, a key 
difference between the proposed and final rule is that we now more 
fully consider the capacity of each species' spatial and demographic 
traits, and habitat heterogeneity, to buffer a species against 
vulnerability to extinction.
    Comment 22: We received a detailed comment letter with supporting 
papers regarding specific effects of ocean acidification on reef-
building corals, such as effects on reef accretion, effects on larvae 
and juvenile corals, and interactive or synergistic effects with other 
environmental variables. For example, the commenter pointed out several 
studies that underscore the potential impact of ocean acidification on 
reef calcification rates, noting that even under the most optimistic 
modeling scenario, 98 percent of reefs would be chemically stressed by 
2050. The commenter also emphasized that corals may have a limited 
ability to adapt to ocean acidification based on an in-situ study of 
two corals in Florida Bay (Okazaki et al., 2013).
    Response: The comment letter and supporting papers support the 
conclusion in the proposed rule that ocean acidification is increasing 
in severity, and likely to continue increasing in severity, within the 
ranges of reef-building corals, resulting in various detrimental 
impacts. This information also underscores the complexity and 
uncertainty associated with the various specific effects of ocean 
acidification, including interactive or synergistic effects with other 
threats, across the ranges of reef-building corals as well as 
predicting adaptive capacity. The information provided by the commenter 
and the supporting papers regarding the specific effects of ocean 
acidification on corals and coral reefs have been incorporated and 
described in more detail in the Threats Evaluation--Ocean Acidification 
sub-section.

Comments on Trophic Effects of Fishing (Medium Importance Threat, ESA 
Factor A)

    Comment 23: One comment provided supplemental information that was 
not included in the proposed rule regarding the role of herbivorous 
fish in terms of building and maintaining reef resilience. The 
commenter stated that ``overfishing also degrades coral reefs, 
particularly by depleting key functional groups, such as herbivores, 
that reduce turf algae on reefs and maintain optimal conditions for 
coral growth and recruitment'' and provided Keller et al. (2009) as a 
reference. Another commenter also described the importance of 
herbivorous functional groups, and stated that limiting or attempting 
to reduce harvest of predatory fish may cause ecological harm by 
unbalancing a healthy trophic chain.
    Response: The proposed rule described the importance of trophic 
interactions which include reducing herbivorous fish species that 
control algal growth, limiting the size structure of fish populations, 
reducing species richness of herbivorous fish, and

[[Page 53863]]

releasing corallivores from predator control. The supplemental 
information provided by public comments supports our conclusion in the 
proposed rule that healthy levels of herbivorous functional groups are 
essential to coral reef ecosystem resilience in light of climate 
change-related threats. Detailed information regarding the trophic 
effects of fishing can be found in the Threats Evaluation--Trophic 
Effects of Fishing sub-section as well as the Inadequacy of Existing 
Regulatory Mechanisms--Reef Resilience sub-section.
    Comment 24: One commenter stated that fish landings have been 
stable for 30 years in St. Thomas, U.S. Virgin Islands, with many 
species increasing in size, indicating that overfishing is not 
occurring in this location or contributing to the status of the 
Caribbean species in that area. The commenter also pointed out numerous 
sources of sediments and nutrients, and coastal development projects in 
the U.S. Virgin Islands as the main contributors to coral reef decline 
rather than overfishing. Other commenters also disagreed that 
overfishing was contributing to coral reef decline in Hawaii and 
highlighted significant increases in tourism and in-water recreational 
activities as local drivers of reef decline in that area.
    Response: Although not explicitly stated in the proposed rule, we 
agree that levels of fishing effort vary throughout the ranges of the 
65 corals under consideration. We did acknowledge that exposure to this 
threat varies throughout the ranges of the proposed species and between 
the Caribbean and Indo-Pacific. In the proposed rule, we also 
recognized that management and regulation of commercial and 
recreational fisheries are inconsistent throughout the coral reef 
world. When evaluating the current and potential threat impacts from 
trophic effects of fishing, we are required to assess this threat 
throughout the entire ranges of the 65 coral species in this final 
listing. We understand that levels and impacts of overfishing differ 
depending on the particular location under evaluation; however, we 
maintain that the trophic effects of fishing represent a medium 
importance threat to the extinction risk of all 65 coral species in 
this final rule.
    Comment 25: One commenter stated that we failed to consider human 
demography in terms of our analysis of fishing impacts to corals. The 
commenter noted that large swaths of area throughout Oceania are being 
depopulated in favor of more metropolitan countries, which reduces the 
level of human impacts to corals, including fishing pressure.
    Response: The issues of human demography and population trends were 
covered explicitly in the SRR and considered in the proposed rule. 
While there may be some areas being depopulated, increased human 
population and consumption of natural resources are root causes for 
increases in fishing (particularly of herbivores) at many locations 
around the globe (Brainard et al., 2011). Data from the World Bank show 
human population abundance and density have increased in all five coral 
reef regions since 1960 (i.e., Indian Ocean, Caribbean, Southeast Asia, 
Pacific, and Middle East), with the greatest human population densities 
and increases in population density in the Southeast Asia and Indian 
Ocean regions. In these regions, current human population densities are 
4-5 times greater than the global average and probably suggest the 
greatest local human-induced effects to corals and coral reefs. In the 
areas in closest proximity to coral reefs, the Southeast Asian, Indian 
Ocean and Middle East regions have the highest densities of people per 
reef area (Burke et al., 2011). However, these data are regional 
averages. We do not dispute that human demography within any of these 
regions may be shifting to higher density in metropolitan areas, 
resulting in a decrease of human disturbance in some portions of these 
regions. The regional trend data suggest increasing risks to corals and 
coral reefs overall (Brainard et al., 2011). However, because we must 
consider the extent to which a particular threat impacts each species 
throughout its entire range, we still maintain that overfishing is a 
medium importance threat to all 65 coral species in this final rule.

Comments on Sedimentation (Low-Medium Importance Threat, ESA Factors A 
and E)

    Comment 26: We received some public comments on sedimentation as a 
threat to the 65 coral species in this final rule. Comments generally 
underscored the importance of sedimentation as a considerable local 
threat to corals and pointed out the potential of sedimentation to 
interact and potentially exacerbate other threats, as well as to reduce 
coral resilience. For example, we received a detailed comment asserting 
that prospects for recovery of certain reef sites in the Caribbean from 
acute episodes of hurricane damage or die-offs from bleaching and 
disease (brought on by ocean warming) are extremely poor without 
sustained recruitment, which may be prevented by sediment preempting 
larval attachment. Further, the commenter identified sedimentation 
(among other local threats) as a local threat with the capability of 
exacerbating bleaching and disease impacts, thereby reducing the 
resilience of corals. One commenter pointed out that mass mortality of 
Acropora palmata at Vega Baja, Puerto Rico, was caused in part by 
sedimentation. Another commenter stated that near shore marine-origin 
sediments have almost completely been replaced by terrestrial sediments 
due to a lack of land use controls, resulting in near total mortality 
of nearshore Acropora stands in the U.S. Virgin Islands. Other 
commenters identified the negative impacts of sedimentation to reefs on 
the Hawaiian Island of Molokai, emphasizing the issue of run-off from 
large rain events in certain areas. In general, these comments 
emphasize the importance of sedimentation as a threat to the 65 coral 
species in this final rule, with some asserting that this threat is as 
important, if not more important, than the higher rated threat of reef 
fishing.
    Response: We acknowledge all of the public comments and information 
we received on the threat of sedimentation to the 65 coral species in 
this final rule. As summarized in the proposed rule, we also recognize 
the possibility for sedimentation to interact with other global and 
local threats and potentially reduce the resiliency of coral reef 
ecosystems and/or impede recovery. In addition to public comments, we 
also collected supplemental scientific information regarding the 
impacts of sedimentation to corals that became available after the 
proposed rule was published. The findings from these studies and more 
detailed information regarding the evaluation of sedimentation as a 
threat to coral reefs can be found in the Threats Evaluation--
Sedimentation sub-section. We also acknowledge the concern that some 
comments expressed regarding the importance of this threat in 
comparison to other local threats. However, for corals in general, we 
maintain that sedimentation is a low-medium threat to the extinction 
risk of the 65 corals in this final rule. Any species-specific 
information we received on sedimentation is included in the Species-
specific Information and Determinations section.

Comments on Nutrients (Low-Medium Importance Threat, ESA Factors A and 
E)

    Comment 27: We received limited public comments on nutrient 
enrichment of nearshore waters (i.e.,

[[Page 53864]]

eutrophication) and its impacts to coral reef ecosystems. Comments 
generally underscored the importance of nutrient enrichment as a 
considerable local threat to corals, and emphasized the potential of 
nutrient enrichment to interact and potentially exacerbate other 
threats, as well as reduce coral reef resiliency. For example, we 
received a detailed comment letter that provided studies regarding the 
impacts of nutrient enrichment to coral species. These studies, which 
became available after the proposed rule was published, provide 
evidence that nutrient enrichment can worsen thermal stress on inshore 
reef communities, and that management actions to reduce coastal 
nutrient enrichment can improve the resistance and resilience of 
vulnerable coastal coral reefs to ocean warming. Another comment 
detailed some of the impacts of nutrients in the U.S. Virgin Islands. 
For example, industrial effluent in St. Croix allegedly impacted 
fisheries in the area to the point where fishermen struggle to sell 
their catch due to perceived contamination. Further, a sewage pumping 
station in another area impacted nursery grounds for spiny lobsters. We 
received other comments regarding the negative impacts of nutrient 
enrichment in various locations in Florida and Hawaii from sewage 
outfalls and other land-based sources of pollution. In general, 
comments emphasized the importance of nutrients as a threat to the 65 
coral species in this final rule, some asserting that this threat is as 
important, if not more, than the higher rated threat of reef fishing.
    Response: In the proposed rule we described the threat nutrient 
enrichment poses to corals. The public comments and supporting papers 
regarding the impacts of nutrients to coral reef ecosystems have been 
considered and incorporated into our assessment, as described in more 
detail in the Threats Evaluation--Nutrients sub-section. We also 
acknowledge the concern that some comments expressed regarding the 
importance of this threat in comparison to other local threats. 
However, for corals in general, we maintain that nutrient enrichment is 
a low-medium threat to the extinction risk of the 65 corals in this 
final rule. Any species-specific information we received on nutrient 
enrichment is included in the Species-specific Information and 
Determinations section.

Comments on Sea-Level Rise (Low-Medium Threat, ESA Factor A)

    Comment 28: We received one public comment that cited the Consensus 
Statement on Climate Change and Coral Reefs (drafted by a working group 
of eminent scientists and endorsed by hundreds of scientists to address 
the topic of climate change impacts on coral reefs; ICRS, 2012) as a 
source of estimates of sea-level rise by the end of this century. 
However, the comment did not expound upon the potential ramifications 
of these estimates. We did not receive any other public comments or 
gather new or supplemental information on the threat of sea-level rise 
to the 65 corals in this final rule.
    Response: Although we received only one public comment on this 
topic, we collected supplemental information regarding the threat of 
sea-level rise to corals as a result of the IPCC's AR5. These findings 
are summarized in the Threats Evaluation--Sea-Level Rise sub-section.

Comments on Predation (Low Threat, ESA Factor C)

    Comment 29: We received very few comments regarding the threat of 
predation to the 65 corals in this final rule. The majority of comments 
we received regarding predation were specific to individual species in 
Guam. For example, we received a detailed comment letter that included 
suggested changes to individual species vulnerability ratings to 
predation, as a result of local crown-of-thorns seastar (Acanthaster 
planci) predation levels. One commenter cautioned us in terms of 
inferring predation vulnerabilities for certain species from genus-
level information. Other comments identified predation as a threat to 
corals, but provided no further information or scientific references.
    Response: We acknowledge all of the public comments and information 
we received on the threat of predation to the 65 coral species in this 
final rule. The extent to which the extinction risk of a coral species 
is impacted by predation is discussed in more detail in the Species-
specific Information and Determinations section, including any 
information we received from specific locations. We also agree that 
inferring susceptibility to threats from genus-level information is not 
always appropriate. However, that particular comment referenced a 
species we deemed Not Warranted for listing under the ESA, and are no 
longer considering. In addition to public comments, we collected 
information regarding the variable effects predation has on certain 
coral species. These studies are detailed and summarized in the Threats 
Evaluation--Predation sub-section. Overall, we maintain that predation 
is a low level threat to the extinction risk of corals in general.

Comments on Collection and Trade (Low Threat, ESA Factor B)

    Comment 30: We received hundreds of comments that strongly 
criticized our characterization of the trade industry as a whole, 
stating that our analysis failed to use current science and/or 
commercial information about the coral trade. Commenters also asserted 
that we did not adequately consider aquaculture and mariculture 
industries as a potential alternative to alleviate pressures from wild 
collection practices. For example, we received a detailed comment 
regarding the mariculture industry in Indonesia, stating that in the 
last five years, the coral trade communities of Indonesia have 
developed coral mariculture with long-term objectives of reducing the 
wild harvest of coral species for the live coral trade. Another comment 
letter provided information from recent papers by Rhyne et al. (2012) 
and Wood et al. (2012) that report declining trade in wild-harvested 
Pacific corals and remarkable growth in the production and trade in 
cultured corals from Pacific countries. Overall, many comments asserted 
that a shift from wild collected corals to cultured corals is occurring 
as a result of increasing aquaculture and mariculture operations both 
within the United States and major source countries such as Indonesia.
    Response: We agree with commenters that the SRR and proposed rule 
did not adequately describe the full scope of the marine ornamental 
trade industry and the contribution of captive culture in terms of 
alleviating pressures from wild collection. We agree that some 
significant progress has been made in terms of shifting from wild 
collection of corals to trade of aquacultured and/or maricultured 
corals as a result of both U.S. domestic production and production of 
corals in major source countries such as Indonesia. In addition to 
public comments we also collected a large amount of supplemental 
information on coral collection and trade. Specifically, we collected 
information about (1) the physical and ecological impacts of wild 
collection of coral colonies and/or fragments from their natural 
habitats; and (2) captive culture (i.e., mariculture and aquaculture) 
including information on operations and the role of home aquaria as it 
relates to trade. All of the public comments and supporting papers have 
been considered and incorporated into our assessment as described in 
more detail in the Threats Evaluation--Collection and Trade sub-
section. However, this information does not change our determination 
that the threat

[[Page 53865]]

is of low importance to the extinction risk of corals, generally.
    Comment 31: We also received numerous comments that strongly 
disagreed with our characterization and conclusion regarding the 
adequacy of regulatory mechanisms within the coral trade industry, 
particularly CITES and other laws in major source countries such as 
Indonesia. Many commenters assert that CITES and various regulations 
provide adequate restrictions and requirements for the ornamental trade 
of coral reef species, such that trade has much less of a negative 
impact on the extinction risk of the 65 coral species than was 
portrayed by the proposed rule and supporting documents. One commenter 
also described Indonesia's development of regulations for their 
mariculture industry that is helping to alleviate wild collection 
pressures.
    Response: In the proposed rule we described that there are some 
protections afforded via CITES and various other national regulations 
in some countries where trade of coral reef species is prevalent. 
However, we agree that our evaluation of trade regulations was 
incomplete. There are numerous challenges in documenting trends in 
trade due to deficiencies of CITES import and export data, and the most 
recent information is conflicting. Some reports state that 98 percent 
of reef-building corals within the aquarium trade are still wild 
collected, with only two percent originating from maricultured sources 
(Thornhill, 2012). In contrast, another report shows that maricultured 
corals accounted for approximately 20 percent of the trade in 2010 
(Wood et al., 2012). Further, adequate tracking of wild and 
maricultured corals along the supply chain from ocean to aquarium is 
extremely difficult, yet necessary for determining the true dimensions 
and impacts of the industry (Cohen et al., 2013). Additionally, the 
level of wild collection of reef-building corals may be underestimated 
due to an undocumented illegal trade and a significant amount of 
mortality along the supply chain from reef to aquarium (Thornhill, 
2012). There are many other issues and discrepancies related to 
assessing the overall impacts of the trade and the adequacy of 
regulations like CITES; however, collection and trade was ultimately 
ranked as a low level threat to corals in general by the BRT and in the 
proposed rule. Further, no one species of coral was determined to be 
threatened or endangered solely due to the effects of the coral trade 
industry, and that is still true for the final determinations in this 
rule. Therefore, while we agree CITES provides some protections for 
corals in the trade industry, we maintain that the threat from 
collection and trade is low and does not dictate the listing status of 
any individual species. In addition to public comments, we collected 
some supplemental information on regulatory mechanisms for the global 
marine ornamental trade industry, including details regarding trade of 
both live and dead corals and other coral reef wildlife.
    In light of the public comments and information we received 
regarding the ornamental trade industry, the Threat Evaluation--
Collection and Trade sub-section discusses the trade and its impacts to 
corals in detail, including information regarding the physical and 
ecological impacts as a result of the collection process, advances in 
aquaculture and mariculture industries, as well as issues and trends in 
trade of both live and dead coral. Any species-specific information we 
received on collection and trade is included in the Species-specific 
Information and Determinations section.

Comments on Inadequacy of Existing Regulatory Mechanisms (ESA Factor D) 
and Conservation Efforts

    Comment 32: We received several comments that critiqued our 
evaluation of local regulatory mechanisms and conservation efforts. 
Some comments asserted that we failed to adequately consider the 
beneficial effects of local management actions and conservation efforts 
with regard to building reef resilience in the face of climate change. 
For example, we received a comment letter that stated a broad consensus 
exists for management to increase marine ecosystem resilience to 
climate change by reducing local anthropogenic stressors and reduction 
of these stressors may boost the ability of species, communities, and 
ecosystems to tolerate climate-related stresses or recover after 
impacts have occurred. Another commenter emphasized the importance of 
local management for increasing coral reef resiliency, including 
management of land-use changes and water quality, as well as utilizing 
coral reef restoration techniques. Overall, these comments disagreed 
with our characterization regarding the effectiveness of local 
regulatory mechanisms and conservation efforts in the face of climate 
change related threats and urged us to consider the concept of reef 
resilience.
    Response: We recognize that certain aspects of local management 
actions and conservation efforts need more explanation than was 
provided in the proposed rule and Management Report (NMFS, 2012b). This 
final rule provides that additional explanation, as summarized here. 
There is an emerging body of literature regarding the concept of reef 
resilience, defined as an ecosystem's capacity to absorb recurrent 
shocks or disturbances and adapt to change without compromising its 
ecological function or structural integrity (Hughes et al., 2010; 
Obura, 2005). Recent evidence suggests that managing local scale 
disturbances for resilience will be crucial to maintaining complex, 
bio-diverse coral reef ecosystems given the predicted widespread 
impacts of climate change related threats (Anthony et al., 2011).
    Therefore, we recognize that effective local laws and regulations 
as well as conservation projects and programs may help reduce impacts 
to corals and coral reefs from threats on an ecosystem level, 
positively affecting the timeframe at which corals may become in danger 
of extinction by providing a protective temporal buffer (i.e., 
resiliency) to individual coral species in the face of climate change 
related threats. Some evidence suggests that local management actions, 
particularly of fisheries (specifically, no-take marine reserves) and 
watersheds, can delay reef loss by at least a decade under ``business-
as-usual'' rises in GHG emissions (Jackson et al., 2014; Kennedy et 
al., 2013; Marshall and Schuttenberg, 2006; Mumby and Steneck, 2011). 
However, many scientists strongly suggest that these local actions be 
combined with a low-carbon economy to prevent further degradation of 
reef structures and associated ecosystems (Kennedy et al., 2013).
    We cannot definitively say whether and to what degree the presence 
of regulations in a particular location is currently conferring 
resilience benefits for any particular species. Overall, we agree that 
local regulatory actions and conservation efforts to reduce threats are 
imperative for resiliency of coral reef ecosystems in the face of 
climate change. However, for purposes of evaluating the inadequacy of 
regulatory mechanisms as well as conservation efforts under the ESA, we 
are unable to definitively establish the current status and 
effectiveness of local regulation of impacts from local threats for any 
particular species in any given location, with the exception of local 
regulatory mechanisms for Acropora palmata and A. cervicornis, which 
were evaluated in detail in the 2005 status review for those species. 
Further, we maintain that global regulations to reduce impacts from 
climate change are inadequate at this time. For more detailed 
information

[[Page 53866]]

about our evaluation of how local regulatory mechanisms relate to 
building coral reef resilience, please refer to the Threats 
Evaluation--Inadequacy of Existing Regulatory Mechanisms sub-section. 
Likewise, for more detailed information about our evaluation of 
conservation efforts please refer to the Conservation Efforts sub-
section.
    Comment 33: We received some comments that disagreed with our 
characterization of local regulatory mechanisms in general, asserting 
that certain local laws are sufficient for protection of corals, thus 
rendering additional protection via the ESA unnecessary. For example, 
we heard from several commenters who believe there are adequate 
regulations to prohibit the damage of reef-building corals, such that 
additional protections from the ESA are redundant. We also received 
comments that disagreed with our characterization of conservation 
efforts. For example, we received a comment that disagreed with our 
conclusion regarding conservation efforts, asserting that coral 
conservation actions already have, and will continue to, contribute to 
coral species recovery. Examples of conservation efforts that were not 
included in the Final Management Report (FMR; NMFS, 2012b) include 
ongoing coral reef restoration projects, specifically in Florida and 
the wider-Caribbean, as well as aquaculture and mariculture efforts 
both internationally (e.g., Indonesia) and within the United States to 
try to alleviate wild collection pressure on coral reef ecosystems. 
Comments urged us to take these efforts into consideration for 
evaluating the status of the 65 corals in this final rule.
    Response: We recognize that certain locations have effective local 
laws, regulations, and programs that address local threats and provide 
for the protection and conservation of coral species. For example, it 
is illegal to collect or harvest reef-building coral species in all 
U.S. states, territories, and commonwealths. Some laws even prohibit 
harming any reef-building coral species through activities such as boat 
groundings and impose penalties and fines for doing so. However, we 
must evaluate whether regulatory mechanisms are inadequate for corals 
across their entire ranges rather than in any one specific location. 
Likewise, our analysis of conservation efforts must also include the 
entirety of the species' ranges, and it must consider whether those 
efforts will result in recovering the species to the point of 
ameliorating threats throughout the species' range to such a degree 
that a species should be listed as threatened rather than endangered or 
even not at all. Therefore, we cannot solely consider whether 
regulations or conservation efforts in the United States or any other 
particular location are sufficient for reducing threats to corals. The 
importance of global climate change-related threats to the extinction 
risk of these corals makes it even more problematic to limit our 
assessment of conservation efforts and the adequacy of regulatory 
mechanisms to individual countries. For these corals, we are required 
to consider the adequacy of regulatory mechanisms for reducing GHG 
emissions and curbing the rate of global climate change.
    For this final rule, we assessed regulatory mechanisms and 
conservation efforts in a more species-specific approach. To better 
capture the full breadth of existing regulatory mechanisms, in addition 
to the individual country descriptions in the Final Management Report, 
we re-characterized and summarized the presence of existing regulatory 
mechanisms throughout all the countries in the range of each individual 
species. The Inadequacy of Threats Evaluation--Existing Regulatory 
Mechanisms sub-section provides more detailed information on that 
range-wide evaluation process, as well as the Species Descriptions for 
the results. For more detailed information about our evaluation of the 
inadequacy of local management actions, please refer to the Threats 
Evaluations--Inadequacy of Existing Regulatory Mechanisms sub-section. 
For more detailed information about our evaluation of conservation 
efforts, please refer to the Conservation Efforts sub-section of this 
rule.
    Comment 34: Several comments identified potential errors, 
omissions, and/or inaccurate characterizations within the Final 
Management Report (NMFS, 2012b). For example, we received a comment 
letter pointing out several omissions and inaccuracies regarding 
Federal management responsibilities for an extensive area of lands and 
waters in the Pacific Ocean. Many other comments provided additional 
laws, regulations, or conservation efforts that were not described in 
the Final Management Report or identified previously during the public 
engagement period during the summer of 2012. For example, one commenter 
requested our inclusion of Guam Public law 24-87 that ensures Guam's 
marine preserves are protected from recreational/commercial activities 
that may prove detrimental to fragile ecosystems. Another commenter 
pointed out that we omitted information regarding certain National 
Wildlife Refuges and National Parks that include coral reefs. We also 
received a public comment letter requesting us to consider information 
regarding Indonesia's Coral Reef Rehabilitation and Management Program 
as a conservation effort.
    Response: We acknowledge that the Final Management Report had some 
minor errors and omissions. However, it should be noted that the Final 
Management Report was not intended to be an exhaustive document; 
rather, it aimed to capture the breadth of existing regulatory 
mechanisms and conservation efforts that may reduce threat impacts to 
corals and coral reefs. Due to the immense number of regulatory 
mechanisms that exist throughout the entire ranges of the 65 coral 
species (i.e., 84 countries), the Management Report was not intended to 
identify every individual law and regulation that may have an effect on 
corals or their threats in every country within the species' ranges. 
However, any additional laws and regulations that were brought to our 
attention through the public comments were noted and considered in the 
analysis of inadequacy of existing regulatory mechanisms presented in 
this final rule under the Threats Evaluation--Inadequacy of Existing 
Regulatory Mechanisms sub-section.

Comments on Risk Analyses

    Comment 35: We received many comments regarding the composition of 
the BRT. Some comments disagreed with the selection of BRT members, 
asserting that because all seven members of the BRT were Federal 
employees, non-Federal coral biologists with expertise in the field 
within specific regions (e.g., Hawaii) were overlooked, thus casting 
doubt on the qualifications of the BRT members and the results of the 
status review. One comment suggested that the BRT member votes should 
have been weighted to reflect their level of expertise in the different 
types of corals undergoing review. Another comment stated that it would 
not be possible for certain members of the BRT to act in a neutral or 
unbiased manner because they are strong proponents of establishing 
Marine Monuments, sanctuaries, and MPAs for the protection of coral 
reef systems throughout the U.S. Pacific Islands. Yet another comment 
stated there was no independent verification from experts who did not 
have a stake in the Federal ESA listing processes.
    Response: According to agency guidance, members of the BRT should 
have expertise in the particular species'

[[Page 53867]]

biology, population dynamics or ecology, or other relevant disciplines 
(e.g., ocean/environmental/climate processes, analytical techniques, 
population genetics, extinction risk, or pertinent threats). 
Additionally, NMFS must also consider team composition in light of the 
Federal Advisory Committee Act (FACA). Generally, any committee or 
group established for the purpose of providing consensus advice or 
recommendations to a Federal agency is subject to the procedural 
requirements of FACA. Biological Review Teams are subject to FACA 
because their assessments constitute group advice upon which NMFS may 
base its determinations as to whether to list species as endangered or 
threatened under the ESA. Based on the requirements of FACA, the team 
must therefore be composed of Federal officials and employees, and 
specific classes of state employees, unless specifically exempted. As 
such, the coral BRT was composed of seven Federal scientists from 
NMFS's Pacific Islands, Northwest, and Southeast Fisheries Science 
Centers and the U.S. Geological Survey and National Park Service. The 
members of the BRT are a diverse group of scientists with expertise in 
coral biology, coral ecology, coral taxonomy, physical oceanography, 
global climate change, and coral population dynamics. Additionally, the 
BRT consulted with numerous non-Federal scientists and subject matter 
experts during the status review, and had their work peer reviewed, to 
ensure the best available information was utilized in the SRR. These 
subject matter experts are detailed in the Acknowledgements of the SRR. 
Last, we provided extraordinary opportunities for non-Federal 
scientists to provide their expertise prior to the publication of the 
proposed rule, including two scientific workshops held in the summer of 
2012. All information received was considered in the proposed rule.
    Comment 36: We received numerous criticisms regarding the 
evaluation methods used by the BRT. Many comments criticized the 
Critical Risk Threshold voting method used by the BRT for developing 
extinction risk values for the 82 corals within the proposed rule. Some 
comments asserted that the voting process relied on subjective opinion 
rather than scientific facts, while other comments stated that the 
anonymous scoring system by the BRT could not truly be anonymous. 
Still, other comments pointed out critical errors or flaws in the BRT's 
methods. For example, one comment stated that ranking each coral 
species relative to the rankings of other coral species does not inform 
NMFS of the risk status of an individual coral species. Another comment 
stated the Critical Risk Threshold graphs have an inappropriate and 
misleading quantitative horizontal axis, which suggests higher threat 
levels than estimated by the BRT. A couple of comments questioned the 
assignment of levels of confidence in the outcomes of the BRT voting 
process given the lack of information on which those outcomes were 
based, noting there was not a high degree of certainty between the 
experts.
    Response: The voting methods used by the BRT are consistent with 
previous agency listing determinations that utilized similar structured 
decision making techniques. This approach is typically used when 
quantitative modeling of extinction risk is not a viable option due to 
a lack of precise quantitative population data. The BRT's voting relied 
upon professional interpretation of the best available scientific 
information at the time, including qualitative assessments. This 
approach allowed the BRT to explicitly address various ranges of 
uncertainty within their voting. We also emphasize that the 
determinations in the proposed rule did not solely rely on information 
within the SRR and the voting outcomes of the BRT. As described 
previously in the proposed rule and throughout this final rule, 
numerous sources of information were considered and incorporated in the 
listing determination process, as described in explicit detail in the 
Risk Analyses and Species-specific Information and Determinations 
sections. Additionally, the ESA does not require quantitative precision 
when estimating extinction risk and determining whether a species 
warrants listing as threatened or endangered under the ESA. Rather, the 
decision must be reasonable and based solely on the best scientific and 
commercial information available at the time of the decision, even in 
light of considerable uncertainty.
    Comment 37: We received several comments that criticized how the 
proposed rule and supporting documents inferred species' 
characteristics based on genus-level information (i.e., the proposed 
rule assumed that information for other species in the genus applied to 
the proposed species in that genus). A few comments stated that the BRT 
only considered threats to the taxonomic class and therefore it 
conducted no individual species threat analysis for any of the 
candidate coral species. Most comments stated that genus-level info on 
response to threats, abundance, and other characteristics were 
improperly extrapolated to species because there are numerous examples 
in the literature in which ecological or physiological traits are not 
consistent across species within a genus.
    Response: In the proposed rule, we relied on higher taxonomic level 
(i.e., genus or family) information for threats susceptibilities when 
species-specific susceptibilities were not available. We acknowledge 
that there is intra-genus or intra-family variability in response to 
threats in many cases. In response to criticism of how the proposed 
rule and supporting documents inferred species' characteristics based 
on genus-level information, this final rule does not automatically 
assume that genus-level information for other species in the genus 
applies to the proposed species in that genus. Rather, a careful 
analysis of genus-level information is incorporated into the Species-
specific Information and Determination sections below for each of the 
21 genera in which the 65 species belong. That is, as a preface to the 
Species-specific Information and Determinations for species in a genus, 
this final rule includes a description of the available information for 
other species in the genus that are not part of this final rule, and an 
analysis of the degree of applicability of that information to the 
species included in this final rule. Further, in no case in this final 
rule do we extrapolate from family-level information.
    Comment 38: We received multiple comments criticizing the 
definition of ``foreseeable future'' in the proposed rule and 
supporting documents out to the year 2100 because it is too far into 
the future. One comment stated that climate change projections beyond 
50 years have a high degree of uncertainty and may be impacted by 
numerous unforeseen and unpredictable circumstances, and thus 
identifying the foreseeable future as out to the year 2100 is not 
appropriate. Another comment stated that our use of 2100 for the 
foreseeable future is contrary to previous decisions made by FWS and 
NMFS, and there have been no breakthroughs in climate modeling to 
justify our new position on the reliability of long-term climate 
modeling.
    Response: Consistent with our practice for all species listing 
determinations, we established that the appropriate period of time 
corresponding to the foreseeable future is a function of the particular 
type of threats, the life-history characteristics, and the specific 
habitat requirements for the coral species under consideration.

[[Page 53868]]

The timeframe established for the foreseeable future considered the 
time necessary to provide for the conservation and recovery of each 
threatened species and the ecosystems upon which they depend. It was 
also a function of the reliability of available data regarding the 
identified threats and extends only as far as the data allow for making 
reasonable predictions about the species' response to those threats. In 
the proposed rule, we explained that our choice of the year 2100 as the 
``foreseeable future'' for analysis of global climate change was based 
on AR4's use of 2100 as the end-point for most of its global climate 
change models. Similarly, most of AR5's WGI models also use 2100 as the 
end-point (some models go beyond 2100) and AR5's WGI reinforces our 
original rationale for defining the foreseeable future as the period of 
time from the present to the year 2100. For global climate change 
threats, there is strong support for considering the foreseeable future 
as the period from the present to 2100 in AR5's WGI and its cited 
literature (IPCC, 2013). However, we agree that the foreseeable future 
for purposes of other threats to the species and the species' responses 
thereto does not necessarily extend out to 2100. Therefore, in this 
final rule, we clarify that 2100 simply marks the outer temporal bounds 
for consideration of climate change-related threats, and does not frame 
our analysis across all threats or our ultimate listing determinations. 
Further discussion of the foreseeable future is presented in the 
Foreseeable Future subsections of the Threats Evaluation and Risk 
Analysis sections below.
    Comment 39: There were many comments on the quantity and quality of 
information used to make listing decisions for the candidate coral 
species in the proposed rule. Several comments stated that the present 
biological data do not support the proposed listings. They stated that 
the available science was insufficient and not compelling enough to 
demonstrate the need to make a decision under the ESA. A few comments 
criticized the use of IUCN data as a surrogate for ``true scientific 
data'' on species distribution and abundance. Many comments stated that 
useful information was available, especially on coral trade and 
mariculture, but the BRT did not use it, which led to serious errors in 
the SRR. The study by Rhyne et al. (2012) was given as an example. 
Other comments stated that there was little data regarding individual 
species' population numbers and trends, so NMFS did not conduct the 
species-specific analyses required under the ESA. In general, the 
commenters indicated that the voting process by the BRT seemed very 
subjective, with the results coming from the individual scientists' 
perception of extinction rather than solid scientific data.
    Response: The proposed rule and the SRR did conduct a species by 
species analysis of extinction risk for each of the candidate corals. 
However, in the proposed rule, the presentation of the information on 
which we based our determinations may have been unclear because of our 
use of the Determination Tool as an organizational mechanism to present 
the enormous amount of data. In response to criticism of the lack of 
sufficient species-specific information in the proposed rule and 
supporting documents on distribution, abundance, threat 
susceptibilities, and other information, this final rule clarifies and 
explains how the information relating to the taxonomic, field 
identification, distribution, abundance, life history, threat 
susceptibilities, and management information for each of the 65 coral 
species were evaluated in reaching the final listing determinations. 
The presentation of the information includes the information on which 
the proposed rule was based, information submitted by public comments, 
and information we gathered after the proposed rule published. The 
information was also analyzed in an integrated, non-formulaic framework 
instead of in a linear, formulaic framework as it was in the 
Determination Tool. The resulting information provides the basis for 
the 65 listing determinations in this final rule. In addition, while 
IUCN listings were used by the petitioner as one criterion for 
selecting coral species in the original 2009 petition, and IUCN maps 
were used in the 2011 SRR, no IUCN data or information is used in this 
final rule because it does not represent the current best available 
species-specific information. To explain more clearly the changes from 
the proposed to the final listings, we included an Overview of Methods 
and Key Changes Applied in Final Determination Framework sub-section 
within the Risk Analyses section to illustrate how all of the available 
information was considered for each species and how it contributed to 
each species' listing status. As a result, the 65 species-specific 
determinations below are based on the best available species-specific 
information and improves upon the proposed rule.
    Comment 40: We received a couple of comments disagreeing with the 
characterization of the level of extinction risk inherent for a species 
due to its occurrence in the Caribbean. One comment stated that the 
BRT's determination that the entire Caribbean is sufficiently limited 
in geographic scale to be a factor that increases the extinction risk 
of all corals in the Caribbean is at odds with genetic data. The 
commenter provided references (Baums et al., 2005b; Baums et al., 
2006a; Murdoch and Aronson, 1999; Vollmer and Palumbi, 2007) that show 
that, while it is clear that regional-scale processes such as bleaching 
and disease are acting on all these reefs simultaneously, no two reefs 
or areas respond the same to these disturbances. Another comment 
asserted that no threat to Caribbean Acropora is imminent, and 
therefore endangered listings are not supportable for these species.
    Response: Geographic distribution is one of many factors we must 
evaluate to determine a species' status. We agree with commenters that 
an inherent increase in extinction risk solely due to occurrence in the 
Caribbean is not accurate; rather, the ratings in the Determination 
Tool regarding basin occupancy were an inadvertent function of 
comparing the Caribbean basin to the Indo-Pacific basin. That is, the 
automatic increase in extinction risk for species occurring in the 
smaller, more disturbed Caribbean was only relative in comparison to 
species occurring in the larger, less disturbed Indo-Pacific. In light 
of public comments, we determined that absolute range size in both the 
Caribbean and Indo-pacific was inadvertently under-estimated in the 
proposed rule. As a result, we now give consideration to geographic 
distribution in terms of absolute size rather than relative size in 
both the Caribbean and Indo-Pacific. We still maintain that the 
Caribbean is a highly disturbed basin that has experienced loss of 
resilience; however, the reconsideration of absolute distribution 
represents one piece of a more holistic approach to linking each 
species' characteristics to each species' status. The implications of 
occurrence in the Caribbean and more detailed descriptions of 
geographic ranges and how they may affect extinction risk are now 
provided in more detail for all species individually in the Risk 
Analyses and Species-specific Information and Determinations sections 
below.
    We also explicitly incorporated consideration of regional and local 
variability in response to threats. We have also endeavored to provide 
a clearer discussion of how we assessed the vulnerability of each coral 
species, not just the Caribbean Acropora, to the major threats. The 
evolution of the

[[Page 53869]]

Determination Tool into a more comprehensive Determination Framework is 
described in the Risk Analyses section of this final rule below.

Comments on the Determination Tool

    Comment 41: Commenters criticized that the Determination Tool 
equated species' characteristics to outcomes without adequate 
rationale. For example, one commenter stated that the Determination 
Tool suffers from a lack of transparency because we did not provide any 
information regarding how the rating values in the Determination Tool 
were assigned, who made the determinations, what their expertise was, 
or on what basis the decisions were made.
    Several comments stated that the Determination Tool's decision 
points and resulting outcomes depended on species-specific information 
that was not available. For example, one commenter asserted that there 
is little to no experimental data provided in the proposed rule 
documentation to support the ratings used in the Determination Tool. 
Another commenter noted that there is almost no information on many of 
the species' abundances, geographic ranges, trends or recruitment 
rates, and that the ratings for these were based solely on qualitative 
opinion. Similarly, another commenter used ocean acidification as an 
example, stating that due to the large degree of uncertainty regarding 
the impacts of ocean acidification on coral species it is difficult to 
quantify the level of risk ocean acidification poses to the species. 
The commenter concluded by stating that assigning levels of ocean 
acidification-associated risk within the Determination Tool is a 
difficult proposition. Another commenter deemed the Determination Tool 
analysis and results arbitrary. The commenter stated that the analysis 
and results of the Determination Tool were based on the same faulty 
assumptions, extrapolations, assessments, and approximations of the 
seven BRT members and were based on very little, if any, species-
specific spatial distribution or abundance data for a number of the 
proposed species.
    Commenters claimed the Determination Tool was flawed and equated 
species' characteristics to listing outcomes too conservatively, 
especially for proposed endangered species. We received a detailed 
comment letter that outlined several criticisms of the Determination 
Tool and its four elements with regard to species outcomes. The 
Determination Tool was labeled faulty because it was deemed inherently 
biased towards listing. The commenter criticized that the first element 
in the Determination Tool was just a re-hash of the BRT's highly 
subjective ranking of threats and vulnerabilities. The commenter also 
asserted that nowhere in the four elements of the Determination Tool is 
there a discussion of imminence or a discussion of why we considered a 
species that meets any of the four factors to be ``on the brink'' of 
extinction. The commenter asserted that we not only failed to adhere to 
the legal standard of endangered, but we did so on extremely poor 
evidence. Further, the commenter criticized the results of the BRT 
voting as well as the Determination Tool for ranking each of the coral 
species' in a relative fashion, and as a result, asserted that our 
approach in determining extinction risk for each species was flawed.
    Several comments pointed out additional perceived flaws in the 
Determination Tool. For example, one commenter stated that the 
Determination Tool dismissed the potential benefits of management 
efforts. Another commenter noted that the Determination Tool did not 
incorporate or consider projections of adaptation potential over the 
foreseeable future (i.e., 2100). More specifically, the commenter 
asserted that the Determination Tool did not sufficiently consider the 
ability of corals to migrate (i.e., undergo range expansion/shift) and 
adapt to changing conditions, especially when local stressors are well 
managed. One commenter also suggested that the Determination Tool 
conflicted with the SRR (e.g., by giving too much weight to 
distribution when the range maps that the BRT relied upon were not 
precise). Similarly, commenters also criticized us for overemphasizing 
the importance of qualitative rankings for species' abundance (e.g., 
common, uncommon, rare) in the Determination Tool, stating that a coral 
species' rarity does not necessarily correspond to its vulnerability to 
extinction.
    Response: We acknowledge that several aspects of the process by 
which we produced our determinations in the proposed rule were not 
described or explained clearly enough. The Determination Tool in the 
proposed rule was a central aspect of a larger overall framework for 
making our decisions, as it organized and standardized our presentation 
of the risk factors, but we acknowledge that the larger determination 
framework was not sufficiently explained in the proposed rule. This 
lack of a clear explanation led to an overemphasis on and 
misunderstanding of the Determination Tool, which was one component of 
the determination framework. To better explain how the Determination 
Tool assessed risk and derived listing statuses we conclude that, as 
some public comments suggested, the Determination Tool was too linear 
and deterministic. We describe our final determination framework in 
greater detail in the Risk Analyses--Final Determination Framework sub-
section below, and utilize a more holistic approach in considering all 
of the available information for each species. As described in that 
section, the non-formulaic approach used in this final rule, is more 
species-specific, and allowed us to address the concern that sufficient 
species-specific information was not available.
    In summary, the Final Determination Framework in this final rule is 
composed of seven elements. The first element is describing the 
statutory standards for corals. The second, third, fourth, and fifth 
elements are identifying and analyzing all the appropriate species-
specific and general characteristics that influence extinction risk for 
a coral species. The sixth element is relating a species' 
characteristics to a particular extinction risk at appropriate spatial 
and temporal scales. The seventh element is explicitly stating how each 
species' extinction risk meets the statutory listing definitions as 
applied to corals, resulting in an ultimate listing status. As a last 
consideration, we determine if any conservation efforts are abating the 
threats to the species such that it changes the individual species' 
listing status (i.e., an endangered species' extinction risk is reduced 
such that it is threatened or that a threatened species is not 
warranted). This method of implementing our Final Determination 
Framework for every species individually is intended to be more 
transparent, by showing how complete use is made of available 
information to reach individual listing decisions.
    We believe that there is still significant value in applying a 
standardized framework to each of the species to ensure consistency 
throughout the 65 individual determinations, but now do that in a 
narrative fashion in which there are no recipes or formulas for 
endangered, threatened, and not warranted species. This approach allows 
for the consideration of the system as a whole (i.e., synergistically 
evaluating each species' demography, spatial characteristics, threat 
susceptibilities, and current and future environmental conditions 
independently of the other species), leading us to species-specific

[[Page 53870]]

conclusions about vulnerability to extinction.
    In response to the criticism that the Determination Tool did not 
appropriately evaluate the imminence of danger of extinction in 
proposing to list corals as endangered, in this final rule we more 
fully explain the biological characteristics and distinctions between 
endangered and threatened corals, and corals not warranting listing 
under the ESA, and relate each species' particular characteristics to 
one of those classifications. These characteristics and the 
distinctions between them as they relate to the three ESA 
classifications are explained in the Statutory Standards sub-section of 
the Risk Analyses section.

Comments on Significant Portion of Its Range (SPOIR)

    Comment 42: We received one comment regarding the identification of 
a significant portion of its range to support not warranted 
determinations for the proposed coral species found in Hawaii. The 
commenter asserted that the species of corals proposed for listing in 
Hawaii are abundant, relatively healthy, and relatively insulated from 
impacts of the primary identified threats. As a result, the species 
will presumably persist in Hawaii, despite more immediate threats in 
other portions of their ranges, ultimately preventing the species from 
going extinct. Thus, the commenter argues that a significant portion of 
its range should be identified for these species, rendering the species 
not warranted for listing.
    Response: The commenter misunderstands the function of the SPOIR 
analysis. As discussed in the Statutory Standard sub-section below, a 
SPOIR analysis is performed to ensure that a species that has been 
found not to be endangered or threatened based on the range-wide 
analysis is still considered for listing if any portions of its range 
meet the criteria of the SPOIR Final Policy. Therefore, the function of 
a SPOIR is not to avoid a listing but to still consider a listing. In 
any case, as described in the Risk Analyses section below, the results 
of our analysis of SPOIR are unchanged from the proposed rule, after 
considering all comments and supplemental information. At this time, no 
SPOIR is determinable for any of the proposed species. Thus, our 
analysis of each species at the range-wide level is determinative, and 
no portions of the range require further examination. Nevertheless, we 
agree with the general underlying premise of the comment, which is that 
if a species has significant areas of refugia or diversity of habitat, 
those are factors that provide additional buffering against extinction 
risk. We have incorporated that consideration in the final rule through 
our Final Determination Framework and species-specific evaluations.

Comments on Listing Determinations

    Comment 43: We received numerous comments referring to the actual 
listing determinations of the 82 candidate coral species in the 
proposed rule. Many of those comments referred to specific coral 
species and to specific aspects of those species determinations. Those 
comments are discussed in detail in the Caribbean Species: Listing 
Determinations, Indo-Pacific Species: Listing Determinations, and 
Reclassification of Acropora palmata and A. cervicornis comment 
response sections below. The other comments regarding listing 
determinations centered on the lack of species-specific information on 
which the species determinations were based. Some comments were very 
skeptical that the assumptions being made from limited scientific 
information on individual coral species, which the proposed rule 
recognized, could justify the listing proposals. These commenters 
asserted that further studies and surveys should be performed to gather 
factual and relevant data on the status of the coral species, which 
could potentially change the assumptions used to make the listing 
determinations. Some comments specifically stated that a much better 
understanding of the global distribution and abundance of the species, 
including developing better species distribution maps, is critical to 
making an informed listing decision. Yet other comments stated that the 
proposed rule did not take into account the variability of response to 
threats that corals can have based on species, location, habitat type, 
and other factors that define an individual coral. Other comments 
suggested that NMFS was using global climate predictions as a 
substitute measure for species population and distribution information 
for listing purposes.
    Response: We recognize that species-specific information was fairly 
limited for many of the proposed species. However, since the proposed 
rule was published, we have received or collected information for all 
species, including supplemental distribution and abundance information 
for 63 of the 65 species in this final rule as a result of the data 
collection effort by Veron (2014). In addition to updating all of the 
general information regarding coral reef biology, ecology, demography, 
and threat susceptibilities, we also incorporated this supplemental 
information in our discussions in the individual species-specific 
listing determinations in that section of this final rule. In light of 
the supplemental species-specific information, and the change to a more 
holistic and species-specific determination framework, we considered 
the ability of each species' distribution and abundance traits to 
affect vulnerability to extinction in the context of the statutory 
definitions of threatened and endangered for each species. In most 
cases, this consideration led to changes in listing status from the 
proposed rule. These species-specific assessments consider all of the 
public comments and available information for each species and provide 
a detailed description of what is and is not known for each species, 
including susceptibilities to all identified threats and vulnerability 
to extinction
    Comment 44: We received several letters alerting us to an extensive 
ongoing effort by coral expert, Dr. J.E.N. ``Charlie'' Veron, to gather 
previously unpublished information for all reef-building corals of the 
world, including the 65 corals in this final rule. The resulting report 
(Veron, 2014) updates species-specific information on semi-quantitative 
(i.e., survey data from 2,984 individual sites) and qualitative 
population abundance estimates (i.e., Veron's subjective estimates 
covering a full range of habitats and most ecoregions the author has 
worked in), geographic distribution, principle habitat, and 
identification issues. Comments stated that given the lack of species-
specific information on quantitative abundances and geographic 
distribution for most of our Indo-Pacific species, this effort proves 
extremely relevant to our listing decisions within this final rule.
    Response: We agree with comments that information from Veron (2014) 
supplemented the existing species-specific information relied on in the 
proposed rule and that this information is relevant to the 
determinations made in this final rule. Thus, the supplemental 
information received in the report (Veron, 2014) was made available to 
the public on NOAA's Web site, and has been incorporated into the 
Species-specific Information and Determinations section for the 63 
species covered in the report, Veron (2014) does not cover non-
scleractinian corals in his report, and thus did not provide 
information for the Millepora species in this final rule).

[[Page 53871]]

Comments on Alternatives To Listing Under the ESA

    Comment 45: We received several comments that suggested 
alternatives to ESA listing such as Candidate Conservation Agreements 
(CCAs), adding the proposed corals to the Species of Concern list, and 
extending the time period in which to make a determination to allow for 
the gathering of additional scientific data. One commenter suggested 
using CCAs, citing lack of species-specific information and other 
concerns as justification. Comments also asserted that because NOAA has 
no authority under the ESA to protect corals from climate change, CCAs 
could provide the same conservation benefits as ESA listings.
    Response: While we acknowledge that CCAs provide conservation value 
for candidate species, no such agreements exist for any of the proposed 
species. Therefore, we are unable to determine a species' status on the 
basis of the conservation provided by a CCA. Further, in the 
Conservation Efforts section we determined that there are no 
conservation efforts currently or planned in the future that are 
expected to improve the overall status of any of the 65 coral species 
in this final rule, such that the additional protections provided by 
the ESA are not warranted.
    We also considered the potential for utilizing the Species of 
Concern designation, which was suggested in lieu of ESA listings due to 
a lack of species-specific information and taxonomic uncertainty. This 
designation can be used if we decide a species is not warranted for 
listing under the ESA because we are unable to confidently assess the 
level of extinction risk, even qualitatively. Ultimately, based on the 
best available scientific information, we concluded that all 65 corals 
within this final rule are determinable species under the ESA. We also 
concluded that we have enough information to qualitatively assess the 
level of extinction risk and make listing determinations for most of 
the 65 species in this final rule with some degree of confidence. The 
species that are determined to be not warranted for listing due to a 
lack of sufficient information to assess their status are clearly 
described as such in the individual species determinations. Those 
species may be considered for inclusion on the Species of Concern list 
after this listing rule becomes final.
    Extending the time period in which to make final species 
determinations in order to collect more scientific data is not 
permissible under the ESA. We are required to use the best scientific 
and commercial data available within the applicable statutory 
timeframes for responding to petitions, as the basis for our final 
determinations.
    Comment 46: We received comments that criticized our proposed 
determinations due to their assumed inconsistency with other recent 
agency decisions, such as the Not Warranted bumphead parrotfish 12-
month finding, and the negative Alaska deep-sea corals 90-day finding. 
Comments cited a lack of adequate species-specific information and 
taxonomic uncertainty as justification for the previous not warranted 
and negative determinations for bumphead parrotfish and Alaskan corals, 
and claimed that the proposed rule for the 68 reef-building corals 
suffers from the same level of uncertainty. Comments thus concluded 
that NOAA's decision to propose 68 reef-building corals for listing 
under the ESA is inconsistent with previous agency decisions and that 
there is a lack of adequate species-specific information to proceed 
with final listings.
    Response: We respond to each petition based on the information 
presented within that petition and, if we conduct a status review, on 
the best scientific and commercial information available for each 
petitioned species at the time. We disagree that this final rule for 65 
reef-building corals is inconsistent with our previous Not Warranted 
12-month finding for the bumphead parrotfish. Primary threats to 
bumphead parrotfish, a coral reef-associated fish, were determined to 
be adult harvest and juvenile habitat loss. As a result of a thorough 
status review, the bumphead parrotfish received a Not Warranted 
determination largely due to its current abundance, life history, 
existing local management in the form of spear fishing regulations, and 
protections for mangrove habitat. Overall, the differences between 
bumphead parrotfish and the reef-building corals in this final rule are 
vast; however, we have complied with the requirements set forth under 
the ESA in each case.
    Likewise, we disagree that this final rule is inconsistent with the 
negative 90-day finding for 44 Alaska deep-sea corals. The Alaska deep-
sea coral species are non-reef building and exhibit many different 
characteristics than shallow-water tropical corals, which have been 
comparatively well researched. The Alaska corals were petitioned due to 
climate change related threats, as well as physical threats from 
commercial fisheries. NOAA considered these factors, but found that 
there are no empirical studies that have shown harmful effects of 
climate change related threats to these deep-sea corals or to similar 
corals in the area. Additionally, ocean acidification research cited in 
the petition was conducted on mostly tropical, reef-building corals 
that are very different from deep-sea corals; no inference could be 
made about the potential effects to the status of deep-sea corals from 
this information. Finally, there have been large swaths of fishing 
ground closures in Alaska since 2005 and NOAA determined that these 
closures were sufficient for protecting deep-water corals from bottom-
contact fishing activities. Overall, differences between the Alaska 
deep-water corals and the reef-building corals in this final rule are 
vast; however, we have complied with the requirements set forth under 
the ESA in each case.

Comments on Caribbean Species: Listing Determinations

    Comment 47: We received some comments that expressed disagreement 
with our proposed threatened determinations for some Caribbean species. 
For example, one comment disagreed with our proposed threatened listing 
of Dichocoenia stokesi, citing the following as justification: Large 
population numbers (even after the White Plague Type II epidemic), 
broad distribution among multiple habitat types (especially hard-bottom 
habitats), high relative abundance among all corals in the region, and 
the presently low prevalence of White Plague Type II. Another comment 
stated that D. stokesi is among the most common species in Florida, and 
that population estimates approached 100 million colonies in 2005, with 
no apparent downward trend. In addition, we received comments about the 
proposed threatened determination for Agaricia lamarcki. Comments 
argued that due to potentially larger populations not yet surveyed in 
deeper waters, the threatened status for A. lamarcki is not warranted. 
Many comments disagreed with the proposed endangered determinations for 
the Orbicella (formerly Montastraea) annularis complex (i.e., O. 
annularis, O. faveolata, and O. franksi). One comment provided 
information from van Woesik et al. (2012) as justification for listing 
O. annularis complex as threatened rather than endangered. Other 
comments submitted a technical report (Miller et al., 2013) from the 
Nova Southeastern University on population abundance estimates and 
trends for the Caribbean coral species in the Florida Keys, in 
opposition to all proposed endangered listing determinations, including 
the proposed endangered determinations for the Orbicella species as 
well as

[[Page 53872]]

Dendrogyra cylindrus and Mycetophyllia ferox. Miller et al. (2013) 
provided recommended changes to the proposed listing statuses for each 
of the proposed Caribbean species using their population and 
distribution estimates as support. We received other comment letters 
that also noted the large population abundances of several of the 
Caribbean species, despite some local declines (i.e., O. annularis and 
O. faveolata). One comment also noted that for five of the Caribbean 
species (i.e., O. franksi, D. cylindrus, M. ferox, D. stokesi, and A. 
lamarcki) there is a complete lack of population data to support ESA 
listings. We also received information regarding the ecology of O. 
annularis and O. faveolata in opposition to their proposed endangered 
determinations, but supporting threatened listings. One comment argued 
that the total population number estimates of these two species are 
very large and that, in light of their broad depth ranges and multi-
habitat distributions, these species are relatively resistant (in 
ecologic time) to extinction. Accordingly, the comment asserted that 
these species' potential listing is contrary to their ecology, 
especially in light of their remaining substantial population numbers 
both in Florida and throughout their range.
    Response: Since the proposed rule was published, we received and 
collected supplemental information for all the Caribbean species, 
including updated distribution and abundance information. Therefore, we 
updated and expanded our individual species-specific descriptions in 
the Species-specific Information and Determinations section for all 65 
reef-building corals within this final rule. These species-specific 
assessments consider the public comments and available information for 
each species, and explain what is and is not known for each species, 
including susceptibilities to the identified threats and overall 
vulnerability to extinction. Further, as described in earlier comment 
responses, we now more fully consider the ability of abundance, 
distribution and habitat heterogeneity to affect vulnerability to 
extinction in the context of the statutory definitions of threatened 
and endangered as applied to corals. The reconsiderations of the 
spatial and demographic factors contributed to changes in all the 
Caribbean species' statuses in this final rule. Thus, as described in 
detail in the Species-specific Information and Determinations section, 
based on the public comments, best available information, and the Final 
Determination Framework, we are revising our proposal to list O. 
annularis, O. faveolata, O. franksi, D. cylindrus, and M. ferox as 
endangered species. Our final determination for these species is to 
list them as threatened species. We have determined D. stokesi and A. 
lamarcki do not warrant listing.

Comments on Indo-Pacific Species: Listing Determinations

    Comment 48: We received several comments regarding our proposed 
threatened and endangered determinations for various Indo-Pacific 
species. Several comments disagreed with our proposed threatened 
determinations for the Hawaiian Montipora clades (M. dilitata/
flabellata/turgescens and M. patula/verrilli). As described in more 
detail below, comments disagreed with the status of these clades and 
suggested they be assessed individually rather than lumped into groups 
(see Comment 49 below for more details). Taxonomic uncertainty as it 
relates to the Genus Montipora and the decision to lump these two 
groups of species is addressed in more detail in the response to 
comments on taxonomic uncertainty (Comment 3 above). Comments also 
asserted that the Montipora clades not only have significantly large 
geographic ranges, but also include some of the most common coral 
species in Hawaii, thus rendering these clades not warranted for 
threatened listing. We received many other comments that disagreed with 
the proposed threatened determinations for a number of the Indo-Pacific 
coral species, but we did not receive any additional substantive 
information or data for consideration of those arguments.
    One commenter provided information regarding the proposed 
endangered status of Pocillopora elegans in the Eastern Pacific. 
Evidence from southwestern Nicaragua suggests that P. elegans has 
undergone extensive mortality, with only a few fragmented and small 
colonies persisting. The data provided, while limited, supports a wider 
body of evidence suggesting particular vulnerability of P. elegans in 
the Eastern Pacific Ocean. However, as described above in Comments on 
Taxonomic Uncertainty in Reef-building Corals, new information on 
Pocillopora species has resulted in our determination that P. elegans 
is not determinable under the ESA.
    The main argument against our proposed endangered determinations 
for Indo-Pacific species is a lack of adequate species-specific 
information to support an endangered status. For example, one comment 
letter noted the percentage of references used in the SRR that provided 
species-specific information for each of the proposed endangered 
species (e.g., only two percent, 5.9 percent and 9.4 percent of the 
references used in the SRR provided species-specific information for 
Acropora rudis, Acropora lokani, and Acropora jacquelineae, 
respectively). We also received comments regarding the proposed 
endangered determinations for various Acropora species, particularly A. 
lokani and A. jacquelineae. For example, one comment emphasized the 
lack of adequate data for the proposed endangered determination of A. 
jacquelineae, citing questionable taxonomic status and lack of density 
estimates and distribution information. Likewise, another comment 
criticized the proposed endangered determination for A. lokani, stating 
that there is virtually no published information available for this 
species. Another comment letter recommended threatened designations for 
A. jacquelineae, A. lokani, and A. rudis rather than endangered, based 
on van Woesik et al. (2012), and stated that Euphyllia paradivisa 
absolutely does not warrant endangered status. We received other 
comments in disagreement with our proposed endangered determinations, 
but they did not include any other substantive information or data to 
consider.
    Response: We recognize that species-specific information was 
limited for many of the Indo-Pacific species. Since the proposed rule 
was published, however, we have received or collected supplemental 
information for several species, including updated distribution and 
abundance information for 63 of the 65 species in this final rule as a 
result of the data collection effort by Veron (2014). As a result, we 
substantially updated and expanded our individual species-specific 
descriptions in the Species-specific Information and Determinations 
section for all 65 reef-building corals within this final rule. These 
species-specific assessments consider all of the public comments and 
available information for each species, and provide a detailed 
description of what is and is not known for each species, including 
vulnerabilities to all identified threats.
    Comment 49: We received some comments that provided species-
specific information for various Indo-Pacific species that is being 
applied in this final rule. The species-specific information we 
received predominantly relates to relative abundance and geographic 
distributions. We specifically received comments on abundance for the 
following Indo-

[[Page 53873]]

Pacific species: Acropora aspera, Porites nigrescens, Acropora diversa, 
and Isopora cuneata. We specifically received comments on distribution 
for the following Indo-Pacific species: Alveopora allingi, Acropora 
palmerae, Acropora paniculata, Acropora jacquelineae, Acropora rudis, 
Euphyllia paradivisa, Acanthastrea brevis, Acanthastrea ishigakiensis, 
Acanthastrea regularis, Acropora globiceps, Acropora lokani, Acropora 
striata, Alveopora fenestrata, Alveopora verilliana, Astreopora 
cucullata, Barabattoia laddi, Euphyllia paraancora, Millepora tuberosa, 
Pavona diffluens, Pocillopora danae, Acropora verweyi, and the 
Montipora clades that are discussed in more detail below. We received 
several detailed comment letters that provided species-specific 
information regarding the Hawaiian Montipora clades (i.e., Montipora 
dilatata/flabellata/turgescens and Montipora patula/verrilli). Several 
of the comments provided references to journal articles or other 
reports as new species-specific information. Some of those references 
were already available to NMFS and some constituted supplemental 
information we did not consider in the proposed rule. We received three 
comments specific to genetics of Indo-Pacific species specifically 
referring to Pavona species at mesophotic depths and to Pocillopora 
species. Species-specific comments regarding taxonomy were specific to 
Acropora acuminata, Acropora paniculata, and Acropora polystoma. 
Comments with species-specific information on threat vulnerabilities 
applied to Acropora aculeus, Acropora aspera, Acropora paniculata, 
Acropora polystoma, Montipora patula, Montipora flabellata, Pocillopora 
elegans, Porites horizontalata, and Seriatopora aculeata.
    Response: Overall, most of the supplemental information we received 
for the Indo-Pacific species was specific to certain geographic 
locations; however, we must evaluate the status of the species 
throughout the entirety of their ranges. As described in earlier 
comment responses, we now more fully consider the ability of spatial 
and demographic traits, as well as the heterogeneous habitats occupied 
by all of the Indo-Pacific species, to affect vulnerability to 
extinction in the context of the statutory definitions of threatened 
and endangered for each species. For many of the Indo-Pacific species, 
their geographic ranges include waters between the east coast of Africa 
and French Polynesia. As described in detail in the Species-specific 
Information and Determinations section, based on the Final 
Determination Framework and supplemental information, we are 
maintaining our proposals to list Acropora globiceps, Acropora 
pharaonis, Acropora retusa, Acropora speciosa, Acropora tenella, 
Isopora crateriformis, Montipora australiensis, Pavona diffluens, 
Porites napopora, and Seriatopora aculeata as threatened in this final 
rule. Five Indo-Pacific coral species determinations changed from 
endangered in the proposed rule to threatened in the final rule: 
Acropora jacquelineae, Acropora lokani, Acropora rudis, Anacropora 
spinosa, and Euphyllia paradivisa. Forty Indo-Pacific coral species' 
determinations changed from threatened in the proposed rule to not 
warranted in the final rule: Acanthastrea brevis, Acanthastrea 
hemprichii, Acanthastrea ishigakiensis, Acanthastrea regularis, 
Acropora aculeus, Acropora acuminata, Acropora aspera, Acropora 
dendrum, Acropora donei, Acropora horrida, Acropora listeri, Acropora 
microclados, Acropora palmerae, Acropora paniculata, Acropora 
polystoma, Acropora striata, Acropora vaughani, Acropora verweyi, 
Alveopora allingi, Alveopora fenestrata, Alveopora verrilliana, 
Anacropora puertogalerae, Astreopora cucullata, Barabattoia laddi, 
Caulastrea echinulata, Euphyllia cristata, Euphyllia paraancora, 
Isopora cuneata, Millepora tuberosa, Montipora angulata, Montipora 
calcarea, Montipora caliculata, Montipora dilatata/flabellata/
turgescens, Montipora lobulata, Montipora patula/verrilli, Pachyseris 
rugosa, Pectinia alcicornis, Physogyra lichtensteini, Porites 
horizontalata, and Porites nigrescens. Finally, Millepora foveolata 
changed from endangered in the proposed rule to not warranted in the 
final rule.
    Last, as described in Comment 2, three coral species determinations 
changed from endangered or threatened in the proposed rule to not 
determinable in the final rule: Pocillopora elegans (eastern Pacific) 
warranted listing as endangered in the proposed rule but was considered 
not determinable in the final rule, and Pocillopora danae and 
Pocillopora elegans (Indo-Pacific) warranted listing as threatened in 
the proposed rule but were considered not determinable in the final 
rule.

Comments on Reclassification of Acropora palmata and Acropora 
cervicornis

    Comment 50: Several comments disagreed with our proposal to 
reclassify the Caribbean species A. cervicornis and A. palmata from 
threatened to endangered. Most comments agreed with the current status 
of threatened for the Caribbean acroporid species. Many comments cited 
increasing abundances, recovering populations, and significant advances 
in active restoration projects as justification for not reclassifying 
them as endangered. One comment opposed the proposed reclassification, 
citing population numbers (Miller et al., 2013), genetic diversity 
(Hemond and Vollmer, 2010), forward-looking population models and 
extinction models based on paleontological data (van Woesik et al., 
2012), and a better understanding of the causes of and resistance to 
mortality (Kline and Vollmer, 2011; Vollmer and Kline, 2008) as 
justification. Comments also stated that there has been no significant 
change in the population status of the acroporids since their initial 
listing in 2006, and populations are relatively stable and recovering 
in some areas. One commenter also emphasized that A. cervicornis in 
particular does not warrant endangered listing status due to its 
presence throughout its entire biogeographical range, population 
expansion northward in south Florida, and its ability to still 
reproduce sexually. One commenter asserted that reclassifying the 
Caribbean Acropora species to endangered is not warranted because the 
threats to these species are not imminent. Additionally, many comments 
cited the growing number of successful restoration projects throughout 
southeast Florida and the Caribbean (Hollarsmith et al., 2012; Johnson 
et al., 2011; Young et al., 2012) that continue to aid in conservation 
of acroporids and help recover genetically viable populations. Overall, 
comments suggest the Caribbean acroporids should remain threatened 
under the ESA, and do not warrant reclassification to endangered 
status. However, we did receive one comment letter in support of the 
reclassifications for the Caribbean acroporids.
    Response: As described previously, we have revised and provided a 
clearer explanation of our decision-making framework to further 
strengthen our final listing determinations. As with all other species 
in this final rule, we updated all of the general information regarding 
coral reef biology, ecology, demography, and threat susceptibilities 
relevant to the Caribbean acroporids, and thus we substantially updated 
and expanded our individual species-specific descriptions for these 
species in the Species-specific Information and Determinations section. 
Further, as previously described in earlier comment responses, we more 
fully consider in

[[Page 53874]]

this final rule the ability of spatial and demographic traits, as well 
as habitat heterogeneity, to affect vulnerability of the Caribbean 
acroporids to extinction in the context of the statutory definitions of 
threatened and endangered for corals.
    We also carefully considered the significant progress of active 
restoration projects in the state of Florida and the wider-Caribbean. 
We agree that these efforts confer conservation and potential recovery 
benefits for the species; however, these efforts, to date, are very 
limited in scale compared to the species ranges and should not be 
considered a panacea for conserving and recovering the Caribbean 
acroporids. The Conservation Efforts section of this rule provides more 
information on active coral reef restoration efforts. As described in 
detail in the Species-specific Information and Determinations section, 
based on the Final Determination Framework and supplemental 
information, we are changing our proposal to reclassify A. palmata and 
A. cervicornis as endangered species. Acropora palmata and A. 
cervicornis will remain listed as threatened species.

Comments on Effects of Listing

    Comment 51: We received several comments that described potential 
negative effects that could result from ESA coral listings. These 
include regulatory burdens in the form of permit applications and other 
various paperwork, consultations and biological opinions, postponement 
of in-water maintenance activities, and increased costs associated with 
harbor improvement projects. We also received numerous comments 
expressing concern about impacts to cultural practices as a result of 
listing, including native artists' livelihoods, reef access by 
indigenous peoples, fishing, lime production, customary navigation and 
seafaring, and specifically native Hawaiian recreational and cultural 
practices, and the cultural needs and practices of American Samoa. One 
comment expressed concern that reclassifying A. palmata and A. 
cervicornis from threatened to endangered will impede ongoing 
restoration and recovery efforts. We received one comment encouraging 
NMFS to make sure we have adequate staff to carry out the additional 
workload associated with ESA Section 7 consultations for any coral 
species that are listed in this final rule.
    Response: The ESA explicitly restricts the factors that can be 
considered in listing decisions. Listing decisions can be based solely 
on the best scientific and commercial data available, after conducting 
a status review and taking conservation measures into account. 
Therefore, comments relevant to the proposed listing include those 
comments that provide additional substantive information regarding 
whether a species is in danger of extinction or likely to become so in 
the foreseeable future (e.g., the biology, status, and/or threats to 
the species, evaluation methodologies, effectiveness of conservation 
measures, accuracy and comprehensiveness of best available information, 
etc.). We are unable to consider other types of comments in a listing 
determination (e.g., socio-economic or policy impacts). However, after 
we implement the final listings, we will work with our stakeholders and 
affected entities to reduce the impact of the listings while still 
providing for the conservation of the listed corals.

Comments on Critical Habitat

    Comment 52: We received three comments related to critical habitat. 
One commenter offered to provide information to assist in the economic 
analysis required for critical habitat designation. A second commenter 
proposed the use of NOAA benthic habitat maps to define areas of 
critical habitat for listed corals and requested reconsideration of 
designated critical habitat for Acropora palmata and Acropora 
cervicornis. A third commenter requested to be consulted during 
critical habitat designation to ensure the operation of their 
facilities would not be affected.
    Response: The comments summarized above do not provide substantive 
information to help inform the final species determinations. NMFS is 
required to designate critical habitat at the time of final rule 
publication, unless we determine that critical habitat is 
undeterminable at that time. Below, we discuss our determination that 
critical habitat is not currently determinable for the species being 
newly listed through this final rule. Designation of critical habitat 
will occur via a separate rule-making process once this final rule is 
published, which will include opportunities for public participation 
and input. As such, the comments described above are noted but are not 
responded to further in this final rule.

Comments on ESA Section 9 Take Prohibitions

    Comment 53: We received 12 comments specific to ESA 4(d) rule-
making, which is discussed in the Section 9 Take Prohibitions section 
of the proposed rule. Eight of these comments requested or suggested 
exemptions from Section 9 take prohibitions for specific activities 
that should be included in a 4(d) rule issued for threatened species 
listed in this final rule. Two comments recommended that lawful 
emissions of GHG should be included as an exception in any future 4(d) 
rule. Two other comments said the opposite, stating that NMFS should 
not consider GHG emissions in the context of the ESA.
    Response: The comments described above did not provide substantive 
information to help inform the final listing determinations for the 65 
coral species. NMFS is not required to issue a 4(d) rule for threatened 
species in conjunction with a final ESA listing. We will do so only if 
we determine it is necessary and advisable for the conservation of 
threatened species. If we make that finding for threatened species 
listed in this final rule, issuance of a 4(d) rule is a separate rule-
making process that will include specific opportunities for public 
input. As such, the comments above are noted but not responded to 
further in this final rule.

Comments on Identification of Those Activities That Would Constitute a 
Violation of Section 9 of the ESA

    Comment 54: We received numerous comments regarding concerns over 
the definition of ``take'' for corals under the ESA. Comments 
questioned how we would define ``take'' if corals are listed, 
considering their unique biological and ecological characteristics 
(i.e., corals are colonial and clonal organisms). One commenter pointed 
out a lack of certainty regarding the threshold of ``take'' for coral 
larvae. Another commenter thought it was unclear how people would know 
if they are ``taking'' a listed coral and expressed concern about the 
ability to conduct cultural practices. A third commenter stated that, 
in the example of corals, the stated goals of the ESA are at odds with 
the best plan for the recovery of any coral species.
    Response: We agree that defining ``take'' of corals under the ESA 
is both unique and challenging, because of the biology of reef-building 
corals. As described below under Corals and Coral Reefs--Individual 
Delineation, these species are both colonial (i.e., capable of creating 
colonies from multiple genetically-identical polyps) and clonal (i.e., 
capable of asexual reproduction to create genetic duplicates). The ESA 
take prohibitions only apply to endangered species immediately upon 
listing. No species in this final rule are being listed as endangered; 
therefore, we do not define activities that may result in take in this 
final rule, because take is not

[[Page 53875]]

automatically prohibited for threatened species. Should we deem it 
necessary and advisable that extending any of the ESA section 9 
prohibitions, including take prohibitions, is necessary for the 
conservation of any of the newly-list threatened coral, we will do so 
in a subsequent rule-making.

Comments on Policies on Role of Peer Review

    Comment 55: We received two comments that criticized NMFS for not 
conducting peer review on the proposed rule. One commenter stated the 
following: ``The Department of Commerce issued guidelines to comply 
with the OMB mandate, publishing the final Guidelines for Ensuring and 
Maximizing the Quality, Objectivity, Utility, and Integrity of 
Disseminated Information in October 2002. As part of the NOAA 
guidelines, the agency must apply a higher standard to `influential 
scientific information' (`ISI'), which is defined as scientific 
information the agency reasonably can determine will have or does have 
a clear and substantial impact on important public policies or private 
sector decisions.' Id. ISI is subject to the more stringent information 
standards in the OMB's Final Information Quality Bulletin for Peer 
Review (``OMB Peer Review Bulletin''), which requires peer review by 
qualified specialists in the relevant field (70 F.R. 2664; January 14, 
2005).''
    Response: The proposed rule itself was not peer reviewed. However, 
the supporting documents that formed the basis for the determinations 
in the proposed rule (e.g., the SRR, FMR) were independently peer 
reviewed by subject matter experts. In addition, much of the 
information we received as a result of the public engagement and public 
comment periods and incorporated into this final rule was independently 
peer reviewed. During the public comment period and subsequent 6-month 
extension solicitation, we received critical review of the information 
on which the proposed rule was based from several coral reef experts. 
As a result, the information used to form the basis of our final 
listing determinations represents the best available scientific and 
commercial information to date on the 65 reef-building coral species 
within this final rule, and that we have complied with all applicable 
policies and guidance on peer review.

Comments Outside of the Scope of the Proposed Rule

    We received numerous public comments in response to the proposed 
rule that are outside the scope of this rulemaking. Below are brief 
explanations to note the comments were received and explain why they 
are not considered relevant to the content of the proposed rule.
    Comment 56: We received several comments regarding concerns over 
potential economic impacts as a result of listing coral species from 
various concerned parties. In addition, we received many comments 
criticizing the proposed rule as an inappropriate use of the ESA to 
protect corals in the face of global climate change. Some comments 
emphasized that the ESA is not designed to regulate GHGs and thus ESA 
listings are not a prudent use of time and resources. Comments also 
cited impacts to cultural practices related to marine resource use in 
opposition of ESA coral listings.
    Response: Due to the statutory requirements of the ESA, comments 
relevant to the proposed listing include those comments that provide 
additional substantive information regarding any facet of the proposed 
rule (e.g., the biology, status, and/or threats to the species, 
evaluation methodologies, accuracy and comprehensiveness of best 
available information, etc.). Comments not relevant to this rule making 
are those comments that are not related to the content of the proposed 
rule and/or comments that we are legally unable to consider in a 
listing determination (e.g., economic impacts). While we are required 
to review and consider all comments, comments on issues outside the 
scope of the proposed rule, such as the comments described above, were 
noted, but are generally not responded to in this final rule.
    Comment 57: Several commenters provided general support for the 
proposed listings but did not provide substantive information or 
specific comments on the content of the proposed rule.
    Response: General support for the proposed action does not 
constitute submission of substantive information regarding any facet of 
the proposed rule. Therefore, these comments were noted but are not 
responded to in this final rule.
    Comment 58: We received three comments pertaining directly to one 
or more of the 16 Not Warranted findings that were issued 
simultaneously with the proposed rule. One commenter questioned why 
some Caribbean species were determined to be Not Warranted while others 
are proposed because threats to all species appear to be the same. 
Another commenter stated that Porites pukoensis should have been 
proposed for listing based solely on the fact that it is endemic to 
Hawaii. A third commenter provided information on Turbinaria 
reniformis' tolerance to threats associated with climate change.
    Response: A Not Warranted finding is a final decision for which 
public comments are not solicited. Therefore, comments on the not 
warranted findings are noted but not considered relevant to the content 
of the proposed rule and are not responded to directly in this final 
rule. We do note, however, that species determinations are based on 
more than just geographic range or existing threats alone and not 
warranted determinations were reached by considering all available 
information on species abundance, range, depth distribution, and threat 
vulnerabilities including susceptibility and exposure, as is described 
in more detail in the not warranted findings.
    As also described in the proposed rule, a threatened coral is 
likely to become an endangered coral within the foreseeable future 
throughout all or a significant portion of its range. For threatened 
species, there is a temporal delay in extinction risk afforded by some 
characteristics of the species, such as broader distribution, larger 
populations, lower vulnerability to the most important threats, and 
better management. Threatened species are less vulnerable than 
endangered species, but still have characteristics that are likely to 
put them at elevated extinction risk within the foreseeable future. For 
each of the 65 species under consideration, we explain how a species' 
characteristics and its ability to provide buffering capacity to the 
identified threats influences its extinction risk over the foreseeable 
future. Some of the 65 species in this final rule meet the definition 
of threatened, as explained in the species sections below.

Basis of Listing Determinations

    The following sections summarize all of the best available 
information on reef-building corals in general, which provides the 
baseline context and foundation for our species-specific listing 
determinations. While this general information illustrates that the 
most important threats are currently increasing in severity, and likely 
to continue increasing further in the foreseeable future, it also 
illustrates that the impacts from these threats, both currently and 
over the foreseeable future, are difficult to interpret and do not 
necessarily correlate to increased vulnerability to extinction due to 
the biological and physical variability and complexity of corals and 
their habitat. Accordingly, our Final Determination Framework and 
species determinations are based upon an analysis of the best

[[Page 53876]]

available species-specific information evaluated within a worsening 
future environment.
    In addition to the comments we received on the proposed rule that 
include new or supplemental information, we have continued to collect 
information that has either emerged since the publication of the 
proposed rule or that was published at the time of the proposed rule, 
but had been inadvertently overlooked. This latter category also 
includes literature cited in the SRR or SIR, but that was further 
examined to provide relevant information. Therefore, we consider 
``supplemental information'' to be that which was not considered at the 
time of the proposed rule that expands upon the themes in the proposed 
rule, but does not fundamentally change a finding from the proposed 
rule. ``New information'' is considered to be that which is novel and 
results in a change to a finding in the proposed rule. To distinguish 
between the information on which the proposed rule was based from new 
or supplemental information, we will only cite the primary literature 
for new or supplemental information. For clarity, we will distinguish 
whether the information was identified via public comment or if we 
gathered it ourselves.
    All the general information on reef-building corals, which provides 
the appropriate context for our species-specific determinations, is 
provided in the Corals and Coral Reefs and Threats Evaluation sections. 
The Risk Analyses section follows and describes our methods and final 
determination framework for making our determinations. Last, we provide 
the individual listing determinations in the Species-specific 
Information and Determinations section, which are based on all of the 
best available information for each coral species.

Corals and Coral Reefs

    This section provides a summary of the best available information 
on the biology and habitat of reef-building corals as it pertains to 
this final rule. First, we briefly summarize the information from the 
proposed rule, which is based on the SRR and SIR. We also address all 
relevant comments received pertaining to the biology and habitats of 
reef-building corals. Further, we provide supplemental information 
relevant to biology and habitat of corals that we gathered during the 
period between the proposed and this final rule. This information 
provides part of the context in which we evaluate the species' status 
and illustrates the unique nature of this evaluation compared to 
typical NMFS' ESA listing determinations (i.e., vertebrates).
    As summarized in the proposed rule, corals are marine invertebrates 
in the phylum Cnidaria that occur as polyps, usually forming colonies 
of many clonal polyps on a calcium carbonate skeleton. The Cnidaria 
include true stony corals (class Anthozoa, order Scleractinia), the 
blue coral (class Anthozoa, order Helioporacea), and fire corals (class 
Hydrozoa, order Milleporina). All 68 proposed species are reef-building 
corals, because they secrete massive calcium carbonate skeletons that 
form the physical structure of coral reefs. Reef-building coral species 
collectively produce coral reefs over time in high-growth conditions, 
but these species also occur in non-reef habitats (i.e., they are reef-
building, but not reef-dependent). There are approximately 800 species 
of reef-building corals in the world.
    Most corals form complex colonies made up of a tissue layer of 
polyps (a column with mouth and tentacles on the upper side) growing on 
top of a calcium carbonate skeleton, which the polyps produce through 
the process of calcification. Millepora fire corals are also reef-
building species, but unlike the stony corals, they have near-
microscopic polyps containing tentacles with stinging cells.

Individual Delineation

    Comment 5 identified the lack of clarity on and complexity of the 
delineation of the ``individual'' with respect to corals and its 
influence in estimating population abundance. We agree that this is a 
complex issue and did not provide sufficient details on how we 
identified what an individual is and how the consideration of this 
issue factored into our estimates of abundances for each of the 
proposed species in the proposed rule. Thus, in this final rule, we 
provide details on how we considered individual delineation in the 
proposed rule and this final rule.
    Reef-building corals are clonal organisms. A single larva will 
develop into a discrete unit (the primary polyp) that then produces 
modular units (i.e., genetically-identical copies of the primary polyp) 
of itself, which are connected seamlessly through tissue and skeleton. 
These modular units may be solitary (e.g., fungiid corals) or colonial. 
Most reef-building coral species are colonial, including all species 
covered in this final rule. Colony growth is achieved mainly through 
the addition of more polyps, and colony growth is indeterminate. The 
colony can continue to exist even if numerous polyps die, or if the 
colony is broken apart or otherwise damaged. The biology of such 
clonal, colonial species creates ambiguity with regard to delineation 
of the individual in reef-building corals, specifically: (1) Polyps 
versus colonies; (2) sexually-produced versus asexually-produced 
colonies; and (3) difficulty determining colony boundaries. Each source 
of ambiguity is addressed below, leading to a conclusion regarding the 
delineation of the ``individual'' for the species covered by this final 
rule, which was not specifically defined in the proposed rule. Though 
not specifically defined, we applied this same concept of the 
individual in the proposed rule.
    The polyp could be considered as the smallest unit of the 
individual for reef-building corals. Each polyp in a coral colony 
consists of a column of tissue with a mouth and tentacles on the upper 
side, growing in a cup-like skeletal structure (the corallite) made of 
calcium carbonate that the polyp produces through calcification. The 
polyps are the building blocks of the colony, and most colony growth 
occurs by increasing the number of polyps and supporting skeleton. 
Polyps carry out the biological functions of feeding, calcification, 
and reproduction. However, because the polyps within a colony are 
modular units, and connected to one another physiologically (i.e., via 
nerve net and gastrovascular cavity, and are the same sex), single 
polyps within a colony are not considered to be individuals for 
purposes of this final rule.
    Alternatively, only colonies originating from sexually-produced 
larvae could be considered as the individual for reef-building corals. 
Colonies are founded by either sexually-produced larvae that settle and 
become the primary polyp of a colony, or asexually-produced fragments 
of pre-existing colonies that break off to form a new colony. Fragments 
from the same colony can fuse back together into the same colony if 
they are close enough to grow together. Fragmentation in branching 
species may lead to a large number of asexually-produced, genetically 
identical colonies, commonly resulting in a population made up of more 
asexually-produced colonies than sexually-produced colonies (Hughes, 
1984). Sexually-produced colonies are important to the population by 
increasing the genetic diversity of the population, and colonies 
originating from asexually-produced fragments do not contribute to the 
effective population (i.e., group of genetically unique individuals). 
Asexual reproduction, though it does not create new genetic 
individuals, is likely the

[[Page 53877]]

more critical mode for some species, especially branching species, 
allowing them to grow, occupy space, and persist between relatively 
rare events of sexual reproduction. Sexually- and asexually-produced 
colonies often cannot be distinguished from one another in the field, 
but are identifiable as an individual, in most cases. Thus, we use the 
concept of the ``physiological colony'' as the entity that can be 
considered an individual. The physiological colony for reef-building 
colonial species is defined here as any colony of the species, whether 
sexually or asexually produced.
    A physiological colony is generally autonomous from other colonies 
of the same species. However, colony morphology, partial colony 
mortality, and other colony growth characteristics (e.g., formation of 
stands or thickets) can complicate the delineation of physiological 
colonies from one another in the field. For example, the overall colony 
morphology of many encrusting species (e.g., some Montipora species) is 
largely dictated by the underlying substrate. In those cases, colony 
shape may not distinguish colonies from one another, and boundaries 
between separate encrusting colonies that have grown together may be 
difficult or impossible to make out visually. Partial mortality of 
colonies, especially larger colonies, can also mask the boundaries 
between colonies, because the algae-encrusted coral skeleton of a 
partially dead colony may appear to delineate two or more colonies. In 
addition, many reef-building coral species occur in stands or thickets 
that may be tens of meters or more in diameter (e.g., some Acropora 
species), possibly consisting of multiple colonies or only one large 
colony, also masking the boundaries between colonies. In each of these 
instances, the actual number of genetically-distinct individuals can 
only be determined through genetic analysis. Those techniques have not 
been established for all coral species and are not feasible to conduct 
for every reef assessment. Therefore, most reef assessments for coral 
abundance also use the concept of the physiological colony as the unit 
for enumerating species.
    Despite the challenges in individual delineation of clonal, 
colonial reef-building corals, this final rule considers the 
``individual'' for each of the proposed species to be the physiological 
colony, as defined above. That is, polyps are not considered 
individuals, but sexually- and asexually-produced colonies are 
considered individuals because they are a type of physiological colony 
and are the unit that can be identified in the field. We acknowledge 
that there are limitations with this definition of the individual, 
including usually-unknown proportions of genetically-distinct 
individuals in a population and the difficulty with the determination 
of physiological colony boundaries. But defining the individual this 
way is the most supportable for this final rule based on the best 
available science. While we did not specifically name the individual as 
the physiological colony in the proposed rule, it is how we considered 
the individual in the proposed rule because the majority of the 
information on abundance is based on the physiological colony which can 
be readily identified and counted in field surveys. Thus, in our 
species determinations we use the physiological colony to inform how we 
estimate abundance of a coral species because that is how field surveys 
estimate coral abundance. Using the physiological colony to estimate 
abundance in the final rule does not change how we estimated abundance 
in the proposed rule, in which we also relied on information that uses 
the physiological colony to report abundance estimates. If we have 
information on the effective population size (i.e., proportion of 
clonality) for a species, that information is also considered.

Taxonomic Uncertainty in Reef-Building Corals

    To determine if the proposed corals meet the ESA definition of a 
species, we had to address issues related to the taxonomic uncertainty 
in corals (e.g., reliance on morphological features rather than genetic 
and genomic science to delineate species) and corals' evolutionary 
history of reticulate processes (i.e., individual lineages showing 
repeated cycles of divergence and convergence via hybridization). To 
address taxonomic uncertainty related to species delineation, except as 
described below where genetic information was available, the proposed 
rule considered the nominal species designation as listed in the 
petition, acknowledging that future research may result in taxonomic 
reclassification of some of the candidate species. Additionally, to 
address complex reticulate processes in corals, the BRT attempted to 
distinguish between a ``good species'' that has a hybrid history--
meaning it may display genetic signatures of interbreeding and back-
crossing in its evolutionary history--and a ``hybrid species'' that is 
composed entirely of hybrid individuals (as in the case of Acropora 
prolifera, discussed in the status review of acroporid corals in the 
Caribbean; Acropora Biological Review Team, 2005). The best available 
information indicates that, while several of the candidate species have 
hybrid histories, there is no evidence to suggest any of them are 
``hybrid species'' (that is, all individuals of a species being F1 
hybrids); thus, they were all considered to meet the definition of a 
``species.''
    Studies elucidating complex taxonomic histories were available for 
several of the genera addressed in the status review, and we were able 
to incorporate those into our species determinations. Thus, while we 
made species determinations for most of the 82 candidate coral species 
on the nominal species included in the petition, we made alternate 
determinations on the proper taxonomic classification for the candidate 
species Montipora dilatata and M. flabellata; Montipora patula and 
Porites pukoensis based on genetic studies. We decided to subsume a 
nominal species (morpho-species) into a larger clade whenever genetic 
studies failed to distinguish between them (e.g., Montipora dilatata, 
M. flabellata, and M. turgescens (not petitioned) and Porites Clade 1 
forma pukoensis). Comment 3 objected to the lumping of the Montipora 
species based solely on one study. However, because the commenter did 
not provide any contrary information and we did not find any new or 
supplemental information suggesting that subsuming the Montipora 
species into a larger clade is incorrect, we are maintaining our 
determination that M. dilitata/M. flabellata/M. turgescens and M. 
patula/M. verrilli are considered species under the ESA.
    In the proposed rule, Pocillopora elegans was split into two 
separate species because the two geographically-distant populations 
have different modes of reproduction. Additionally, the proposed rule 
examined the listing status of P. danae. After consideration of the 
information on taxonomic uncertainty, including from the proposed rule 
and supporting documents, Comment 2, and new information, we have 
determined that these three Pocillopora species (P. elegans (Eastern 
Pacific), P. elegans (Indo-Pacific), and P. danae), are not listable 
entities under the ESA. As explained in the response to Comment 2, new 
information on the three proposed Pocillopora species proposed for 
listing indicates an increasing level of taxonomic uncertainty to the 
point that these three species are not listable entities under the ESA 
at this time. Thus, this final rule considers 65 of the 68 species 
included in the proposed rule. However, even though these

[[Page 53878]]

remaining 65 species are determinable under the ESA, some uncertainty 
regarding taxonomy and certain species identification remains. These 
uncertainties are addressed for each species in the Species-specific 
Information and Determinations sections.
    In addition to these specific examples of species delineation, 
Comment 1 stated that taxonomic uncertainties associated with many 
reef-building coral species are problematic for the ESA listing 
determination process. We acknowledge the clear delineation among 
individuals that characterizes vertebrate species is often absent in 
reef-building coral species. This final rule addresses that ambiguity 
with the general introductions in this sub-section, then by providing 
species-specific information for each species. Therefore, the level of 
taxonomic uncertainty is addressed for each of the species in this 
final rule in the Species-specific Information and Determinations sub-
sections below.

Species Identification

    We received several comments related to the difficulty in coral 
species identification (see Comment 1). In the proposed rule we 
acknowledged the difficulty in identification and how that affected the 
ability to accurately infer abundances for individual species (see 
proposed rule Distribution and Abundance section). However, we did not 
discuss the species identification uncertainty on a species by species 
basis. In this sub-section, we more fully describe the challenge of 
species identification. In the Species-specific Information and 
Determinations section, we address the identification uncertainty for 
each species, and determine if that uncertainly affects the reliability 
of the distribution and abundance information described for each 
species, based on expert analysis (Fenner, 2014b).
    In this final rule ``species identification'' refers to the 
assignment of a given individual to a species based on its appearance 
in the field or lab. In contrast, ``species delineation'' refers to the 
definition of reef-building corals as distinct species based on their 
scientific classification or taxonomy (covered in the previous sub-
section). Many reef-building coral species are difficult to identify 
for many reasons, including: (1) The high biodiversity of reef-building 
corals; (2) the high morphological plasticity in many reef-building 
coral species; and (3) the different methods used for species 
identification. An example of all three factors working together (high 
biodiversity, morphological plasticity, different methods) is provided 
by massive Porites species: Many species occur together in the same 
habitats and locations, morphological plasticity is high for both 
colony shape and corallite structure, and experts disagree about how to 
distinguish the species (Forsman et al., 2009; Veron, 2000).
    Coral species identification is based on the assumption that the 
taxonomy is correct. The high biodiversity, high morphological 
plasticity, and different methodologies create species identification 
problems even when the taxonomy is correct. But if the taxonomy is not 
correct, the species identification problems described here are 
irrelevant because species with a high level of taxonomic uncertainty 
(e.g., the Pocillopora species in this final rule) are not listable 
entities under the ESA. Both the species delineation and species 
identification problems are highly species-specific, and are addressed 
for each species in the Species-specific Information and Determinations 
section.

Reproductive Life History of Reef-Building Corals

    As summarized in the proposed rule, corals use a number of diverse 
reproductive strategies that have been researched extensively; however, 
many individual species' reproductive modes remain poorly described. 
Most coral species use both sexual and asexual propagation. Sexual 
reproduction in corals is primarily through gametogenesis (i.e., 
development of eggs and sperm within the polyps near the base). Some 
coral species have separate sexes (gonochoric), while others are 
hermaphroditic. Strategies for fertilization are either by ``brooding'' 
or ``broadcast spawning'' (i.e., internal or external fertilization, 
respectively). Asexual reproduction in coral species most commonly 
involves fragmentation, where colony pieces or fragments are dislodged 
from larger colonies to establish new colonies, although the budding of 
new polyps within a colony can also be considered asexual reproduction. 
In many species of branching corals, fragmentation is a common and 
sometimes dominant means of propagation.
    Depending on the mode of fertilization, coral larvae (called 
planulae) undergo development either mostly within the mother colony 
(brooders) or outside of the mother colony, adrift in the ocean 
(broadcast spawners). In either mode of larval development, larvae 
presumably experience considerable mortality (up to 90 percent or more) 
from predation or other factors prior to settlement and metamorphosis. 
Such mortality cannot be directly observed, but is inferred from the 
large amount of eggs and sperm spawned versus the much smaller number 
of recruits observed later. Coral larvae are relatively poor swimmers; 
therefore, their dispersal distances largely depend on the duration of 
the pelagic phase and the speed and direction of water currents 
transporting the larvae. The documented maximum larval life span is 244 
days (Montastraea magnistellata), suggesting that the potential for 
long-term dispersal of coral larvae, at least for some species, may be 
substantially greater than previously understood and may partially 
explain the large geographic ranges of many species.
    The spatial and temporal patterns of coral recruitment have been 
studied extensively. Biological and physical factors that have been 
shown to affect spatial and temporal patterns of coral recruitment 
include substrate availability and community structure, grazing 
pressure, fecundity, mode and timing of reproduction, behavior of 
larvae, hurricane disturbance, physical oceanography, the structure of 
established coral assemblages, and chemical cues. Additionally, factors 
other than dispersal may influence recruitment, and several other 
factors may influence reproductive success and reproductive isolation, 
including external cues, genetic precision, and conspecific signaling.
    In general, on proper stimulation, coral larvae settle and 
metamorphose on appropriate substrates. Some evidence indicates that 
chemical cues from crustose coralline algae, microbial films, and/or 
other reef organisms or acoustic cues from reef environments stimulate 
settlement behaviors. Calcification begins with the forming of the 
basal plate. Buds formed on the initial corallite develop into daughter 
corallites. Once larvae are able to settle onto appropriate hard 
substrate, metabolic energy is diverted to colony growth and 
maintenance. Because newly settled corals barely protrude above the 
substrate, juveniles need to reach a certain size to limit damage or 
mortality from threats such as grazing, sediment burial, and algal 
overgrowth. In some species, it appears that there is virtually no 
limit to colony size beyond structural integrity of the colony 
skeleton, as polyps apparently can bud indefinitely.
    Comment 4 identified the lack of information on coral population 
dynamics and connectivity; however, it did not provide any supplemental 
information, other than for Acropora

[[Page 53879]]

cervicornis, which will be considered in that species' determination. 
Therefore, the section above is a summary of the information on coral 
reproductive life history from the proposed rule as it contributes to 
the extinction risk analyses for the proposed corals. In our species 
determinations, we consider life history characteristics that may 
contribute to extinction risk. For example, species with high 
recruitment rates or fast growth rates may have the ability to more 
quickly recover from disturbances. Additionally, long-lived species 
with large colony size can sustain partial mortality (fission) and 
still have potential for persistence and regrowth. However, detailed 
life history information is not available for all of the species 
considered in this final rule, though it is used when available.

Distribution and Abundance of Reef-Building Corals

    The proposed corals are distributed throughout the wider-Caribbean 
(i.e., the tropical and sub-tropical waters of the Caribbean Sea, 
western Atlantic Ocean, and Gulf of Mexico; herein referred to 
collectively as ``Caribbean''), the Indo-Pacific biogeographic region 
(i.e., the tropical and sub-tropical waters of the Indian Ocean, the 
western and central Pacific Ocean, and the seas connecting the two in 
the general area of Indonesia), and the tropical and sub-tropical 
waters of the eastern Pacific Ocean. In our species determinations, 
spatial and demographic traits inform our evaluation of a species' 
current status and its capacity to respond to changing conditions over 
the foreseeable future. One important demographic trait is absolute 
abundance, which is a function of local density (either quantitative or 
qualitative) and range size. Absolute abundance is more informative 
than a relative description of abundance for corals such as ``rare,'' 
because even a coral species described as ``rare'' may still have 
millions of individual colonies or more (i.e., few individuals per unit 
area spread across a very large area). Similarly, the spatial trait of 
geographic distributions are not considered on a relative scale (i.e., 
narrow, moderate, wide as we did in the proposed rule), but rather 
considered on an absolute scale, which for even the smallest species 
distribution encompasses millions of square miles.
    As described in the Individual Delineation sub-section, determining 
abundance of the proposed corals presents a unique challenge because 
corals are clonal, colonial invertebrates, and colony growth occurs by 
the addition of new polyps. In addition, colonies can exhibit partial 
mortality in which a subset of the polyps in a colony dies, but the 
colony persists. Colonial species present a special challenge in 
determining the appropriate unit to evaluate for status. In addition, 
new coral colonies, particularly in branching species, can be added to 
a population by fragmentation (breakage from an existing colony of a 
branch that reattaches to the substrate and grows) as well as by sexual 
reproduction (see above, and Fig. 2.2.1 in SRR). Fragmentation results 
in multiple, genetically identical colonies (ramets) while sexual 
reproduction results in the creation of new genetically distinct 
individuals (genotypes or genets).
    In the proposed rule, quantitative abundance estimates were 
available for only a few of the candidate species. In the Indo-Pacific, 
many reports and long-term monitoring programs describe coral percent 
cover only to genus level because of the substantial diversity within 
many genera and difficulties in field identification among congeneric 
species. In the Caribbean, most of the candidate species are either too 
few in numbers to document meaningful trends in abundance from 
literature reports (e.g., Dendrogyra cylindrus), or commonly identified 
only to genus (Mycetophyllia and Agaricia spp.), or potentially 
misidentified as another species. At the time of the proposed rule, the 
only comprehensive abundance data in the Caribbean were for the three 
Orbicella species, partially because they historically made up a 
predominant part of live coral cover. Even for these species, the time 
series data are often of very short duration (they were not separated 
as sibling species until the early 1990s and many surveys continue to 
report them as ``Orbicella annularis complex'') and cover a very 
limited portion of the species range (e.g., the time series only 
monitors a sub-section of a single national park). In general, the 
available quantitative abundance data were so limited or compromised 
due to factors such as small survey sample sizes, lack of species-
specific data, etc., that they were considerably less informative for 
evaluating the risk to species than other data, and were therefore 
generally not included as part of the individual species extinction 
risk evaluations.
    Comment 47 provided quantitative abundance estimates from Florida 
for all of the proposed corals in the Caribbean. In addition, we 
gathered supplemental information providing quantitative abundance 
estimates and distribution for individual species in the Caribbean and 
Indo-Pacific. These data are included and described in the individual 
extinction risk assessments for those species in the Species-specific 
Information and Determinations section.
    Unlike quantitative abundance data, qualitative abundance 
characterizations (e.g., rare, common), were available for all species 
(Veron, 2000), and were considered in the proposed rule's individual 
species extinction risk evaluations. These estimates are the subjective 
opinion of the author and are meant to indicate relative abundance 
between the categories. That is, a rare species has fewer individuals 
as compared to an uncommon one, and an uncommon species has fewer 
individuals than a common one. These estimates are also meant to 
describe the author's opinion of the qualitative abundance of the 
species throughout its range, and not necessarily an estimate of the 
abundance at an individual location. Since the proposed rule was 
published, semi-quantitative (i.e., survey data from 2,984 individual 
sites) and updated non-quantitative (i.e., the author's subjective 
estimates covering a full range of habitats and most ecoregions the 
author has worked in) abundance estimates were provided for 63 of the 
65 corals covered in this final rule (Veron, 2014). In addition to the 
semi-quantitative and non-quantitative estimates, Veron (2014) provided 
occupancy of each species within the approximately 150 ecoregions he 
has defined. An ecoregion is defined as an area that is internally 
cohesive (i.e., areas with similar habitats share similar species 
complements), but externally distinct from neighboring regions (http://coral.aims.gov.au/). Ecoregions are widely used in biogeography because 
they incorporate a substantial amount of background knowledge, are a 
good platform for statistical analysis, and allow the pooling and 
comparison of different datasets from the same ecoregion. Ecoregions 
are not equal in size and thus occupancy in the same number of 
ecoregions by two different species does not indicate the same range 
size. Rather, the number of ecoregions occupied is a good indication of 
the diversity of habitats and geographic distribution in which a 
species may be found. These data are included in the individual 
extinction risk assessments for those species in the Species-specific 
Information and Determinations section.
    As previously described in the Individual Delineation section, 
clonal, colonial organisms, such as corals, are vastly different in 
their biology and ecology than vertebrates, which are typically the 
focus of ESA status reviews. Therefore, concepts and terms that are 
typically applied to vertebrates have very distinct meanings when

[[Page 53880]]

applied to corals. A `rare' coral may still have millions of colonies 
as compared to a `rare' vertebrate, which may only have hundreds of 
individuals.

Coral Habitats

    As summarized in the Coral Reefs, Other Coral Habitats, and 
Overview of Candidate Coral Environments section of the proposed rule, 
a ``coral reef'' is a complex three-dimensional structure occurring 
from the surface to approximately 30 to 40 meters of depth resulting 
from the skeletal growth of reef-building corals that provides habitat, 
food, and shelter for numerous marine species. As such, coral reefs 
foster exceptionally high biodiversity and provide the following 
essential functional roles: Primary production and recycling of 
nutrients in relatively nutrient poor (oligotrophic) seas, calcium 
carbonate deposition yielding reef construction, sand production, 
modification of near-field or local water circulation patterns, and 
habitat for secondary production, including fisheries. These functional 
roles yield important ecosystem services in addition to direct economic 
benefits to human societies such as traditional and cultural uses, food 
security, tourism, and potential biomedical compounds. Coral reefs 
protect shorelines, coastal ecosystems, and coastal inhabitants from 
high seas, severe storm surge, and tsunamis.
    The three broad categories of coral reefs are fringing reefs, 
barrier reefs, and atolls. Fringing reefs are mostly close to 
coastlines, and usually have a high component of non-carbonate 
sediment. Barrier reefs are offshore and are composed of wave-resistant 
consolidated limestone. Atolls are usually a wall of reefs partially or 
completely enclosing a central lagoon. There are not sharp differences 
that clearly mark boundaries between reef types. For example, fringing 
reefs gradually become barrier reefs with increasing distance from 
shore. Also, the shape of both barrier reefs and atolls is largely 
determined by the bathymetry of the substratum, producing many 
irregularly shaped reefs that are intermediary between the two types. 
Isolated reefs that do not fit any of these descriptions are referred 
to as platform reefs (Veron, 2000).
    Despite the differences between the reef categories, most fringing 
reefs, barrier reefs, atolls, and platform reefs consist of a reef 
slope, a reef crest, and a back-reef, which in turn are typically 
characterized by distinctive habitats. The reef slope is the seaward 
side of the coral reef between the reef crest and the deep ocean, and 
generally includes upper fore-reefs or upper slopes (approximately 5-10 
to 10-20 m depth), mid-slopes that often occur as terraces or shelves 
(approximately 10-20 to 20-30 m depth), and deep fore-reefs, lower 
slopes, or walls (approximately 30-40 m depth) that transition to 
mesophotic areas (greater than 30-40 m depth). The reef crest 
(approximately 0 to 5-10 m depth) forms the boundary between the reef 
slope and back-reef, and generally includes a consolidated ridge or rim 
where the waves break, and a lower reef crest on the seaward side of 
the algal ridge often made of up of buttresses and surge channels 
(i.e., spur-and-groove structures). The back-reef lies between the reef 
crest and land (or middle of the lagoon, in the case of atolls). The 
back-reef generally includes reef flats (approximately 0 to 1-5 m 
depth) and lagoons (approximately 1-5 to over 30 m depth), interlaced 
with tide pools, channels, patch reefs, and other features. The 
characteristics of these habitat types vary greatly by reef categories, 
locations, latitudes, frequency of disturbance, etc., and there is also 
much habitat variability within each habitat type, together 
constituting the habitat heterogeneity of coral reefs, as described 
further below.
    Fringing reefs occur adjacent to coastlines, and subsequently the 
habitats associated with their reef slopes and back-reefs may be quite 
different than on barrier reefs or atolls. The reef slopes of many 
fringing reefs that are protected from strong wave action (e.g., on 
leeward sides of islands) consist of unconsolidated material sloping 
gently towards deeper water, while those of fringing reefs in more 
exposed areas (e.g. windward sides of islands) are usually more 
consolidated. On many fringing reefs, even on the reef slope, natural 
turbidity and sedimentation may be high due to proximity to land. 
Fringing reefs typically have narrow back-reefs consisting of a reef 
flat abutting the reef crest, and possibly tide pools, channels, or 
small lagoons between the reef flat and shore (Goreau, 1959; Veron, 
2000). Barrier reefs typically form tens to hundreds of kilometers from 
coastlines, their reef slopes are composed of consolidated limestone 
that may plunge steeply to deeper water, and natural turbidity and 
sedimentation are very low due to distance from land. Thus the 
characteristics of their reef slope habitats can be quite different 
than on fringing reefs. Barrier reefs are exposed to very strong wave 
action, and their reef crests can vary from high, consolidated algal 
ridges to unconsolidated shingle ramparts to low and wide indistinct 
crests. In addition, barrier reefs typically have immense back-reefs 
consisting of reef flats abutting the reef crest, and large lagoons 
that may vary from clear and sandy near the reef to turbid and muddy 
near land, and include various features such as patch reefs and islands 
(Maxwell, 1968). Atolls occur in oceanic waters far from land, and may 
be hundreds of kilometers across. Their reef slopes often form vertical 
walls dropping into abyssal waters, and their back-reefs consist of 
large, clear lagoons (Veron, 2000; Wells, 1951). Environmental 
conditions vary greatly between the habitat types found on the reefs 
slopes, reef crests, and back-reefs of the world's coral reefs. In 
addition, much variability also occurs within each habitat type. For 
example, Maxwell (1968) describes six geomorphological types of reef 
crests, and how the different environmental conditions provide ``coral 
zones'' unique to each type of reef crest. The physical diversity of 
coral reef habitat is illustrated by Kuchler (1986), who notes that the 
scientific literature on the GBR alone used over 20 terms for the reef 
slope or its habitats, over 50 terms for the reef crest or its 
habitats, and over 100 terms for the reef flat and lagoon and their 
habitats.
    In conclusion, five main points are important regarding coral 
habitat on coral reefs (as opposed to non-reefal and mesophotic 
habitats) for this final rule: (1) Regardless of reef category, reefs 
generally consist of reef slopes, reef crests, and back-reefs, each of 
which have distinct habitats, but those habitats can be highly variable 
between reef types and locations; (2) spatial variability in coral 
habitat conditions is very high between habitat types, as well as 
within the habitat types described above (i.e., deep fore-reefs, walls, 
mid-slopes, upper reef slopes, lower reef crests, algal ridges, reef 
flats, and lagoons), producing highly variable environmental conditions 
across both large and small spatial scales at any given point in time; 
(3) temporal variability in coral habitat conditions is also very high, 
both cyclically (e.g., from tidal, seasonal, annual, and decadal 
cycles) and episodically (e.g., storms, temperature anomalies, etc.); 
(4) together this spatial and temporal variability in environmental 
conditions across multiple scales produces the very high habitat 
heterogeneity of coral reefs; and (5) while most coral species in this 
final rule are more common in certain reef habitat types, they are 
typically found in many different habitat types.
    Reef-building corals have specific habitat requirements, including 
hard substrate, narrow mean temperature

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range, adequate light, and adequate water flow. These habitat 
requirements most commonly occur on the shallow tropical and 
subtropical coral reefs described above, but also occur in non-reefal 
and mesophotic areas. All of the proposed species require hard 
substrates. Thus, in this final rule, ``non-reefal habitats'' refers to 
hard substrates where reef-building corals can grow, including marginal 
habitats where conditions prevent reef development (e.g., turbid or 
high-latitude or upwelling-influenced areas) and recently available 
habitat (e.g., lava flows). The term ``mesophotic habitats'' refers to 
hard substrates deeper than 30 m. Coral reefs, non-reefal areas, and 
mesophotic areas are not necessarily sharply delineated from one 
another, thus one may gradually blend into another. We anticipate the 
total area of non-reefal and mesophotic habitats is greater than the 
total area of shallow coral reef habitats within the ranges of the 
corals in this final rule.
    Comments 6 and 7 suggested that we did not consider non-reefal 
habitats and mesophotic habitats adequately in our proposed rule. 
However, these comments did not provide any new or supplemental 
information on how to interpret the importance of these habitats in our 
extinction risk analysis. Comment 7 includes two studies that provide 
supplemental information on the extent of mesophotic reefs. In addition 
to the public comment received on the diversity and complexity of coral 
reef habitats, supplemental information has become available on non-
reefal and mesophotic habitats since the publication of the proposed 
rule. The following sub-sections on non-reefal and mesophotic habitats 
are intended to illustrate the diversity of reef-building coral 
habitats, but are not intended to provide an exhaustive list of them.
    Non-reefal habitats include marginal habitats (Perry and Larcombe, 
2003), as well as newly available natural habitats such as the hard 
substrates created by lava flows (Grigg and Maragos, 1974), tsunamis 
(scoured bedrock or transported boulders (Goto et al., 2010)), or other 
episodic processes. Non-reefal habitats are defined as areas where 
environmental conditions prevent reef formation but reef-building 
corals are present. Marginal habitats are much more common than newly-
available natural habitats. Marginal habitats are very diverse, as they 
occur where seawater temperatures or light levels are sub-optimal 
(i.e., inadequate for high skeletal growth but still allowing reef-
building corals to survive), and thus include environments that are 
turbid (Blakeway et al., 2013; Browne et al., 2012), very warm (Riegl 
and Purkis, 2012; Riegl et al., 2011), or cold because of high latitude 
(Dalton and Roff, 2013; Lybolt et al., 2011) or upwelling (Alvarado et 
al., 2011; Manzello et al., 2008), and other environments (Couce et 
al., 2012; Done, 1982; Perry and Larcombe, 2003). Some coral species 
can also live on soft substrates, such as Manicina areolata in the 
Caribbean, staghorns (Acropora) that must begin on hard substrate but 
can then grow over soft substrates, and Catalaphyllia jardini, which is 
common in some soft substrates in Australia. Such habitat is not 
necessarily indicative of low-diversity coral assemblages, as shown by 
turbid sites, which have been documented to support over 160 species of 
reef-building corals (Perry and Larcombe, 2003), and fresh lava flows, 
which have been documented to support fully recovered coral communities 
only 20 years after the flow (Grigg and Maragos, 1974). Marginal 
habitats expands the diversity of environmental conditions that can 
support some reef-building corals and therefore may provide refugia 
from some threats affecting shallow coral reef habitat, as described in 
the Spatial and Temporal Refugia sub-section below.
    Since 2012, research on mesophotic habitats has demonstrated that 
many reef-building corals have greater depth distributions than 
previously reported. Twenty-two of the proposed species have been 
reported from mesophotic depths (i.e., 30 m or more) and several more 
reported at 25 m. For other species, their biogeographic ranges may be 
underestimated due to lack of mesophotic exploration. These studies 
demonstrate that some species in shallow coral reef habitats readily 
extend to mesophotic depths if water clarity and temperatures remain 
favorable (Kahng et al., 2014). For example, investigations in American 
Samoa (Bare et al., 2010), the Hawaiian Archipelago (Kahng et al., 
2010; Rooney et al., 2010), and the Mariana Archipelago (Rooney et al., 
2012), have revealed extensive mesophotic coral reef ecosystems. While 
classically considered to be limited to 100 m, mesophotic reefs have 
been observed as deep as 130 m in some of these areas, including at 
depths in excess of 150 m in the Au`au Channel of Hawaii (Blyth-Skyrme 
et al., 2013). Likewise, investigations on Australia's GBR found 
extensive mesophotic habitats both along the continental shelf-edge and 
on submerged reefs inside the lagoon of the GBR, both of which support 
previously unknown communities of reef-building corals (Bridge et al., 
2012a; Bridge and Guinotte, 2013; Bridge et al., 2012b). As noted in 
one of these recent papers, several coral species (including Acropora 
aculeus, A. jacquelineae, and A. tenella) are common and geographically 
widespread in deeper waters (30-60 m; Bridge et al., 2013b). Other 
recent studies in Cura[ccedil]ao (Bongaerts et al., 2013), Bermuda 
(Locke et al., 2013), and Hawaii (Luck et al., 2013) reveal extensive 
mesophotic habitats and reef-building coral communities. These studies 
expand the known potential habitats for reef-building corals, but 
species diversity and abundances have not been well- documented due to 
the relative inaccessibility of these habitats to divers.
    In summary, the magnitude of habitats potentially supporting reef-
building coral species is extremely large, and much larger than the 0.2 
percent of the marine environment provided in the SRR. Globally, some 
reef-building corals can occur in shallow coral reef, non-reefal, and/
or mesophotic habitats. These three types of general habitats combined 
provide the overall physical environment of many species, and 
supplemental information on non-reefal and mesophotic habitats 
indicates that their magnitude is larger than previously understood.

Inter-Basin Comparisons

    As described in the proposed rule, the Caribbean and Indo-Pacific 
basins contrast greatly both in size and in condition. The Caribbean 
basin is geographically small and partially enclosed, has high levels 
of connectivity, and has relatively high human population densities. 
The wider-Caribbean occupies five million square km of water and has 
approximately 55,000 km of coastline, including approximately 5,000 
islands. Shallow coral reefs occupy approximately 25,000 square km 
(including [ap]2,000 square km within U.S. waters), or about 10 percent 
of the total shallow coral reefs of the world. The amount of non-reefal 
and mesophotic habitat that could potentially be occupied by corals in 
the Caribbean is unknown, but is potentially greater than the area of 
shallow coral reefs in the Caribbean.
    The Caribbean region has experienced numerous disturbances to coral 
reef systems throughout recorded human history. Fishing has affected 
Caribbean reefs since before European contact, and continues to be a 
threat. Beginning in the early 1980s, a series of basin-scale 
disturbances has led to altered community states, and a loss of

[[Page 53882]]

resilience (i.e., inability of corals and coral communities to recover 
after a disturbance event). Massive, Caribbean-wide mortality events 
from disease conditions of both the keystone grazing urchin Diadema 
antillarum and the dominant branching coral species Acropora palmata 
and Acropora cervicornis precipitated widespread and dramatic changes 
in reef community structure. None of the three important keystone 
species (Acropora palmata, Acropora cervicornis, and Diadema 
antillarum) have shown much recovery over decadal time scales. In 
addition, continuing coral mortality from periodic acute events such as 
hurricanes, disease outbreaks, and bleaching events from ocean warming 
have added to the poor state of Caribbean coral populations and yielded 
a remnant coral community with increased dominance by weedy brooding 
species, decreased overall coral cover, and increased macroalgal cover. 
Additionally, iron enrichment in the Caribbean may predispose the basin 
to algal growth. Further, coral growth rates in the Caribbean have been 
declining over decades.
    Caribbean-wide meta-analyses suggest that the current combination 
of disturbances, stressful environmental factors such as elevated ocean 
temperatures, nutrients and sediment loads, and reduced observed coral 
reproduction and recruitment have yielded a loss of resilience, even to 
natural disturbances such as hurricanes.
    Coral cover (percentage of reef substrate occupied by live coral) 
across the region has declined from approximately 50 percent in the 
1970s to approximately 10 percent in the early 2000s (i.e., lower 
densities throughout the range, not range contraction), with concurrent 
changes between subregions in overall benthic composition and variation 
in dominant species. However, supplemental information suggests that 
this estimate of coral cover decline in the Caribbean is an 
oversimplification. In the Caribbean, quantitative surveys of a few 
dozen sites from before the early 1980s suggest the regional mean for 
coral cover was 30-40 percent around 1980 (Gardner et al., 2003; 
Schutte et al., 2010). Supplemental information based on more complete 
sampling effort (i.e., meta-analysis of 35,000 quantitative reef 
surveys from 1969 to 2012) indicates higher levels of ``current'' 
percent live coral cover in the Caribbean than described in the 
proposed rule. For example, a recent study found that average coral 
cover throughout the wider-Caribbean declined by 66 percent from an 
overall average of 41 percent between 1969-1983 to 14 percent today, 
slightly higher than the 10 percent reported earlier. The earlier 
reports were based on less thorough sampling of the available data, and 
were also dominated by data from the Florida Keys, U.S. Virgin Islands, 
and Jamaica, which may not be representative of the entire Caribbean 
(Jackson et al. 2014).
    In conclusion, the supplemental information regarding live coral 
cover does not dispute that there has been a long-term overall decline 
in live coral cover in the Caribbean and that those declines are likely 
ongoing and likely to continue in the future as a result of a multitude 
of global and local threats at all spatial scales. These wide-scale 
changes in coral populations and communities have affected habitat 
complexity and may have already reduced overall reef fish abundances. 
These trends are expected to continue. However, as the above 
information illustrates, live coral cover trends are highly variable 
both spatially and temporally, producing patterns on small scales that 
may not be indicative of conditions throughout the basin.
    Ocean basin size and diversity of habitats (e.g., reef-flats, 
forereef, mesophotic, non-reefal), as well as some vast expanses of 
ocean area with only very local, spatially-limited, direct human 
influences, have provided substantial buffering of Indo-Pacific corals 
from many of the threats and declines manifest across the Caribbean. 
The Indo-Pacific (Indian and Pacific Oceans) is enormous and hosts much 
greater coral diversity than the Caribbean region (~700 coral species 
compared with 65 coral species). The Indo-Pacific region encompasses 
the tropical and sub-tropical waters of the Indian Ocean, the western 
and central Pacific Ocean, and the seas connecting the two in the 
general area of Indonesia. This vast region occupies at least 60 
million square km of water (more than ten times larger than the 
Caribbean), and includes 50,000 islands and over 40,000 km of 
continental coastline, spanning approximately 180 degrees of longitude 
and 60 degrees of latitude. There are approximately 240,000 square km 
of shallow coral reefs in this vast region, which is more than 90 
percent of the total coral reefs of the world. In addition, the Indo-
Pacific includes abundant non-reefal habitat, as well as vast but 
scarcely known mesophotic areas that provide coral habitat. The amount 
of non-reefal and mesophotic habitat that could potentially be occupied 
by corals in the Indo-Pacific is unknown, but is likely greater than 
the area of shallow coral reefs in the Indo-Pacific (NMFS, 2012b; SIR 
Section 4.3).
    While the reef communities in the Caribbean may have poor 
resilience, the reefs in the central Pacific (e.g., American Samoa, 
Moorea, Fiji, Palau, and the Northwestern Hawaiian Islands) appear to 
remain much more resilient despite major bleaching events from ocean 
warming, hurricanes, and crown-of-thorns seastar predation outbreaks. 
That is, even though the reefs have experienced significant impacts, 
corals have been able to recover, as described below. Several factors 
likely result in greater resilience in the Indo-Pacific than in the 
Caribbean: (1) The Indo-Pacific is more than 10-fold larger than the 
Caribbean, including many remote areas; (2) the Indo-Pacific has 
approximately 10-fold greater diversity of reef-building coral species 
than the Caribbean; (3) broad-scale Caribbean reef degradation likely 
began earlier than in the Indo-Pacific; (4) iron enrichment in the 
Caribbean may predispose it to algal growth versus lack of broad-scale 
iron enrichment in the Indo-Pacific; (5) there is greater coral cover 
on mesophotic reefs in the Indo-Pacific than in the Caribbean; and (6) 
there is greater resilience to algal phase shifts in the Indo-Pacific 
than in the Caribbean.
    Even given the relatively higher resilience in the Indo-Pacific as 
compared to the Caribbean, one meta-analysis of overall coral status 
throughout the Indo-Pacific indicates that substantial loss of coral 
cover (i.e., lower densities throughout the range, but not range 
contraction) has already occurred in most subregions. As of 2002-2003, 
the Indo-Pacific had an overall average of approximately 20 percent 
live coral cover, down from approximately 50 percent since the 1970s. 
However, supplemental information refines this estimate. Data from 154 
surveys of reefs across the Pacific performed between 1980 and 1982 had 
mean live coral cover of 42.5 percent (Bruno and Selig, 2007). Coral 
cover in the Indian Ocean declined from approximately 40 percent prior 
to the 1998 bleaching event to approximately 22 percent; subsequently, 
mean coral cover increased to approximately 30 percent by 2005 
(Ateweberhan et al., 2011) Live coral cover likely had already declined 
in all regions before 1980, but region-wide quantitative data is 
generally lacking. For example, local surveys before 1980 from several 
parts of the Indo-Pacific documented live coral cover of 50 to 70 
percent (Gomez et al., 1981).
    Unlike the Caribbean, no recent region-wide reports of current, 
overall live coral cover are available for the Indo-Pacific as a whole. 
However,

[[Page 53883]]

recent reports from parts of the region have found current live coral 
cover higher than the 20 percent for the region reported earlier, and 
stable or increasing live coral cover. For example, monitoring data 
collected annually from 47 sites on the GBR from 1995 to 2009 averaged 
29 percent live coral cover. More importantly, this study found no 
evidence of consistent, system-wide decline in coral cover since 1995. 
Instead, fluctuations in coral cover at sub-regional scales (10-100 
km), driven mostly by changes in fast-growing Acroporidae, occurred as 
a result of localized disturbance events and subsequent recovery 
(Osborne et al., 2011). However, another recent study based on 2,258 
surveys of 214 GBR reefs over 1985-2012, showed declines in live coral 
cover from 28 percent to 14 percent, a loss of half of the initial 
coral cover. In the Philippines, a study of 317 sites from 1981 to 2010 
averaged 36 percent live coral cover, and showed an overall increase 
from 29 percent in 1981 to 37 percent in 2010 (Magdaong et al., 2013). 
A study of 366 sites from 1977 to 2005 in the Indian Ocean documented 
large initial decline from approximately 35 percent live coral cover to 
approximately 15 percent at most sites following the 1998 bleaching 
event, followed by partial recovery to approximately 25 percent, and 
then stability of live coral cover (Ateweberhan et al., 2011). 
Likewise, a study in Western Australia from 2005 to 2009, following the 
2005 bleaching event, documented declines to 10 percent live coral 
cover as a result of the event and then subsequent recovery to 30 
percent (Ceccarelli et al., 2011). A study in the Andaman Islands from 
2010 to 2012 following the 2010 bleaching also documented substantial 
recovery of live coral cover (Marimuthu et al., 2012; Osborne et al., 
2011).
    These recent studies illustrate the dynamic nature of live coral 
cover, especially recovery from the 1998 bleaching event. It is likely 
that the overall basin-wide live coral cover in both the Caribbean and 
the Indo-Pacific has declined over decadal and centurial time scales, 
but with fluctuations on shorter time scales and within smaller 
geographic scales. This is significant because coral decline doesn't 
occur in every location at every time scale. Rather, there are periods 
of decline and recovery over shorter time periods in various locations 
throughout the larger basins. This has broad implications when 
analyzing the temporal and spatial elements of a coral species' 
extinction risk.
    Disagreements over the methods of how to measure live coral cover 
have led to different results in studies measuring changes in live 
coral cover over time. For example, one study (Bellwood et al., 2004) 
reported approximately 50 percent declines in live coral cover on GBR 
over the last several decades, but another study disagreed (Sweatman et 
al., 2011), making the case for considerably smaller declines, using a 
different method. Both studies provided detailed support for their 
methods and findings (Hughes et al., 2011; Sweatman and Syms, 2011). 
Studies supporting both results have since been published (De'ath et 
al., 2012; Osborne et al., 2011), and such disagreements illustrate the 
complexity of determining trends in live coral cover.
    In conclusion, the supplemental information regarding live coral 
cover does not dispute that there has been a long-term overall decline 
in live coral cover in both the Caribbean and Indo-Pacific, and that 
those declines are likely ongoing and likely to continue in the future 
due to a multitude of global and local threats at all spatial scales. 
Further, both basins have experienced conditions leading to coral 
mortality and prevention of full recovery; however, the Caribbean has 
been more greatly impacted. While basin-wide averages are useful for 
large-scale comparisons, they do not describe conditions at finer, 
regional scales. For example, decreases in overall live coral cover 
have occurred since 2002 in some areas, such as on the GBR, while 
increases have occurred in other areas, such as in American Samoa. As 
the supplemental information further illustrates, live coral cover 
trends are highly variable both spatially and temporally, producing 
patterns on small scales that can be easily taken out of context. Live 
coral cover trends are complex, dynamic, and highly variable across 
space and time. Thus their interpretation requires the appropriate 
spatial-temporal context (i.e., entire range or each species now and 
through foreseeable future), and an understanding of the various 
physical, biological, and ecological processes at work within coral 
communities and coral reef ecosystems.
    In the proposed rule, we provided a summary of conditions in the 
eastern Pacific to illustrate the contrast to the conditions in Indo-
Pacific and Caribbean. This description was relevant because the range 
of one of our candidate species, Pocillopora elegans (eastern Pacific), 
was restricted to the eastern Pacific. Because we are no longer 
considering the three proposed Pocillopora species in this final rule, 
a detailed description of the eastern Pacific is not necessary.

Spatial and Temporal Refugia

    Comment 7 suggested that certain habitats (e.g., mesophotic) may 
provide refugia for shallow water corals. Therefore, we provide the 
following discussion of temporal and spatial refugia. Some of these 
concepts were discussed in the Threats Evaluation section of the 
proposed rule as they relate to exposure of corals to the various 
threats and how exposure influences extinction risk. The above 
information on coral habitats illustrates the enormous heterogeneity of 
the environments that many of these species inhabit. Each species 
occurs in a patchwork of variable habitat conditions at any given point 
in time, with certain combinations of variables at certain locations 
producing favorable conditions that may provide refugia from threats 
such as ocean warming. Habitat conditions are highly variable over time 
in different ways, including cyclically (e.g., from tidal, seasonal, 
annual, and decadal cycles), episodically (e.g., storms, temperature 
anomalies, etc.), and linearly (e.g., gradual thermal regime changes, 
which will both degrade and improve habitat, depending on location and 
initial conditions). The dynamic nature of reef-building coral habitats 
may provide refugia for some corals from some threats, both spatially 
and temporally (Fine et al., 2013; McClanahan et al., 2011; Riegl and 
Piller, 2003).
    Some habitats have natural features that reduce stress from 
extremely high temperatures or light levels (i.e., the most common 
causes of coral bleaching), which may provide spatial refugia for some 
reef-building coral species from ocean warming and other threats. 
Deeper water may be cooler depending on the amount of mixing, and is 
exposed to less light (i.e., irradiance). Mesophotic habitats are very 
extensive, and recent investigations provide evidence that mesophotic 
habitat functions as refugia for some reef-building corals. A review of 
mesophotic habitat on Australia's GBR concluded that reef-building 
corals in mesophotic habitat are less likely to be affected by warming-
induced bleaching events than their counterparts on nearby shallow 
reefs (Bridge et al., 2012a). Mesophotic habitat may also be important 
for recovery of corals disturbed coral reefs by providing sources of 
propagules to recolonize shallow reefs following disturbances (Bridge 
and Guinotte, 2013). A 37-year record from the eastern Pacific across 
the two most severe El Ni[ntilde]o events on

[[Page 53884]]

record (1982-83 and 1997-98) shows how an exceptionally thermally-
sensitive reef-building fire coral, Millepora intricata, twice survived 
catastrophic bleaching in a deeper water refuge (>11 m depth). During 
both events, M. intricata was extirpated across its range in shallow 
water but showed recovery within several years, while two other fire 
corals without deep-water populations were driven to regional 
extinction (Smith et al., in press).
    The refuge value of mesophotic habitats is limited, however. Only 
about one-quarter of all reef-building coral species occur at 
mesophotic depths (Bongaerts et al., 2012) and only 22 of our proposed 
species. Also, there is limited connectivity between mesophotic and 
shallow coral habitats, at least for some species, suggesting that the 
actual likelihood of mesophotic corals repopulating shallow reef 
habitats is low for those species. For example, genetic connectivity 
between mesophotic and shallow populations is high in Seriatopora 
hystrix on the GBR (van Oppen et al., 2011) and Millepora intricata in 
the eastern Pacific (Smith et al., in press), but low for Montastraea 
cavernosa in the Caribbean (Brazeau et al., 2013).
    Marginal habitats are also extensive, and recent investigations 
provide evidence that marginal habitat also functions as refugia for 
some reef-building corals. Marginal habitats include turbid (Blakeway 
et al., 2013; Browne et al., 2012), very warm (Riegl and Purkis, 2012; 
Riegl et al., 2011), cold (Dalton and Roff, 2013; Lybolt et al., 2011), 
soft substrate, and other environments (Couce et al., 2012; Done, 1982; 
Perry and Larcombe, 2003) with sub-optimal coral growth conditions. A 
study of future coral habitat suitability under ocean warming and 
acidification suggests that marginal habitats may provide important 
refugia for some reef-building corals (Couce et al., 2013b), though not 
all coral species can survive in these habitats. The study found that 
the IPCC AR4's higher emission scenarios are all likely to result in: 
(1) Range expansion at the high-latitude boundaries; (2) no decreased 
suitability in currently marginal eastern Equatorial Pacific locations 
as well as in the Atlantic generally; and (3) severe temperature-driven 
impacts in the western Equatorial Pacific (Coral Triangle) and 
surrounding regions. These findings led to the conclusion that marginal 
habitat is likely to function as a patchwork of refuge habitats for 
some reef-building corals in both the Indo-Pacific and Atlantic as 
ocean warming and acidification increase over the twenty-first century.
    Aside from mesophotic and marginal habitats, other types of 
habitats may provide refuge for reef-building corals from ocean warming 
and other threats. Some of these have long been known to reduce thermal 
stress, such as those habitats with highly-fluctuating conditions, 
strong currents from wind or tides, and shading from frequent cloud 
cover or complex bathymetry, as described in the proposed rule and 
supporting documents. Supplemental information suggests other 
oceanographic features may also provide refuge from ocean warming both 
currently and the foreseeable future, such as: (1) Large-scale 
upwelling in both the Pacific (Karnauskas and Cohen, 2012) and 
Caribbean (Bayraktarov et al., 2012); (2) the similar but smaller-scale 
phenomenon of internal tidal bores that transport cooler, deeper water 
to warmer, shallower areas (Storlazzi et al., 2013); (3) and the wakes 
of relatively cool water left by the passage of tropical cyclones 
(Carrigan, 2012). Most of the refugia described above are with regard 
to ocean warming, but some of these habitat types provide refugia 
potential from ocean acidification, such as highly-fluctuating habitats 
which limit pH minima via tidal flux (Shaw et al., 2012), and from 
disease and sedimentation, such as high-energy habitats which provide 
flushing that reduces conditions conducive to disease and removes 
sediment. Seagrass beds provide beneficial changes in ocean chemistry 
to seawater on adjacent reefs, providing local refugia to ocean 
acidification (Manzello et al., 2012). Depth also provides some refugia 
potential from disease, as most studies show a negative correlation 
between depth and coral disease incidence. However, some studies show 
no such correlation, and disease incidence can be comparable between 
mesophotic and shallow depths (Brandt et al., 2012).
    Thermal regime changes from ocean warming will have opposite 
effects on habitat, depending on location: In locations already near 
the thermal maxima of reef-building corals, warming will degrade 
habitat, but in locations currently too cool for these species, warming 
will improve habitat, if other habitat features conducive to reef 
growth are also present, such as hard substrate and appropriate light 
and water chemistry conditions. Geological evidence from past global 
warming periods shows a pattern of poleward expansion of some reef-
building coral ranges, coupled with decline in equatorial areas 
(Kiessling et al., 2012) and expansion into temperate areas (Woodroffe 
et al., 2010). Predicted ocean warming in the twenty-first century is 
expected to result in a similar pattern of poleward expansion, thus 
newly-colonized areas may provide temporary refugia for some species 
(van Hooidonk et al., 2013b). For example, models suggest that such 
expansion of reef-building corals could occur at the rate of 1-4 km per 
year in Japan (Yara et al., 2011). As temperatures increase to the 
optimal range for reef-building corals in these northerly and southerly 
areas, however, the simultaneous increase in ocean acidification may 
negate the suitability of these areas (van Hooidonk et al., 2014; Yara 
et al., 2012). While it may appear that there is no long-term, large-
scale refugia from both ocean warming and ocean acidification (van 
Hooidonk et al., 2014), on a finer regional and/or reef-scale, there is 
still a large amount of refugia in the form of heterogeneous habitat, 
including mesophotic, non-reefal, and marginal habitats, that provide a 
buffer to corals from threats into the foreseeable future.

Corals and Coral Reefs Conclusion

    The above general information on reef-building coral biology and 
habitat leads to several important overall points that apply both 
currently and over the foreseeable future. With regard to reef-building 
coral biology, first, delineations between individual colonies of the 
same species, and between species, can be highly uncertain, creating 
ambiguity with regard to the status of species--specific sources of 
uncertainty include unclear individual delineations, taxonomic 
uncertainty, and species identification uncertainty. Thus, in our 
species determinations we use the physiological colony to inform how we 
estimate abundance of a coral species because that is how field surveys 
estimate coral abundance. Using the physiological colony to estimate 
abundance in the final rule does not change how we estimated abundance 
in the proposed rule, in which we also relied on information that uses 
the physiological colony to report abundance estimates. If we have new 
or supplemental information on the effective population size (e.g., 
proportion of clonality) for a species, that information is also 
considered. Second, while corals can reproduce both sexually and 
asexually, abundance estimates are based solely on the physical number 
of coral colonies that does not recognize mode of reproduction. 
Dispersal and recruitment patterns are highly variable across space and 
time, leading to complex and poorly understood population dynamics and

[[Page 53885]]

connectivity. In our species determinations, we consider life history 
characteristics that may contribute to extinction risk. For example, 
species with high recruitment rates or fast growth rates may have the 
ability to more quickly recover from disturbances. Additionally, long-
lived species with large colony size can sustain partial mortality 
(fission) and still have potential for persistence and regrowth. Third, 
all species considered in this final rule occur in multiple habitat 
types and have considerable distributions that encompass at least 
thousands of islands and multiple habitat types, which influences 
absolute abundances--the absolute distributions and absolute abundances 
of these species are key components of their vulnerability to 
extinction. Therefore, in our species determinations, the spatial and 
demographic traits of absolute abundance and absolute distribution 
inform our evaluation of a species' current status and its capacity to 
respond to changing conditions over the foreseeable future.
    Additionally, because of variability between species, some 
generalities cannot be assumed to apply equally to each species. 
Therefore, in our species determinations we consider the complex nature 
of coral biology and assume that for all species, responses to threats 
will be variable between individual coral colonies and even between 
different portions of the same colony. The best available species-
specific information for each of the 65 species is provided in the 
Species-specific Information and Determinations sub-sections below.
    With regard to reef-building coral habitat, first, the 
heterogeneity of reef-building coral habitat varies greatly both 
spatially and temporally. That is, the habitat of a given species 
varies spatially (i.e., even the smallest ranges of the species 
included in this final rule encompass thousands of islands and multiple 
habitat types) and temporally (i.e., varies over time in response to 
disturbances and recoveries). Second, some habitat types are 
understudied (e.g., mesophotic and marginal) so data about their 
contribution to the distribution and abundance of individual coral 
species are limited, as well as the possibility of refugia from 
particular threats being underestimated. Third, a diversity of habitats 
likely helps some species capacity to acclimatize and adapt to changing 
conditions, especially extreme habitats. For example, while some 
colonies die during the stressful conditions common to extreme 
habitats, other colonies at the same reef survive and acclimatize, 
potentially leading to adaptation. The magnitude and diversity of reef-
building coral habitats creates high physical heterogeneity across the 
ranges of these species, providing habitat refugia from threats. Some 
of these refuge habitats may already be occupied by the species; others 
could become occupied as their suitability changes, assuming the 
species are able to reproduce and successfully recruit into these 
areas. The habitat heterogeneity and refugia lead to variable micro-
climates at a reef scale that leads to variable responses by reef-
building corals to threats, both spatially and over time, which adds 
complexity to assessing the status of species in a worsening 
environment.
    Overall, in our species determinations, we recognize that the 
exposure and response of a coral species to global threats varies 
spatially and temporally based on variability in the species' habitat 
and distribution. All species considered in this final rule occur in 
multiple habitat types, or reef environments, and have distributions 
that encompass diverse physical environmental conditions that influence 
how that species responds to global threats. As such, the concept of 
heterogeneous habitat influences extinction risk for all species in 
this final rule because each species experiences a wide variety of 
conditions throughout its range which allows for variable responses to 
global and local threats.

Threats Evaluation

    Section 4(a)(1) of the ESA and NMFS' implementing regulations (50 
CFR 424) state that the agency must determine whether a species is 
endangered or threatened because of any one or a combination of five 
factors: (A) Present or threatened destruction, modification, or 
curtailment of habitat or range; (B) overutilization for commercial, 
recreational, scientific, or educational purposes; (C) disease or 
predation; (D) inadequacy of existing regulatory mechanisms; or (E) 
other natural or manmade factors affecting its continued existence. In 
the proposed rule, our evaluation of the five factors was informed by 
the SRR and SIR for factors A-C and E; and the Final Management Report 
for factor D. We identified factors acting directly as stressors to the 
82 coral species (e.g., sedimentation and elevated ocean temperatures) 
as distinct from the sources responsible for those factors (e.g., land 
management practices and climate change) and qualitatively evaluated 
the impact each threat has on the candidate species' extinction risk 
over the foreseeable future.
    The proposed rule qualitatively ranked each threat as high, medium, 
low, or negligible (or combinations of two; e.g., ``low-medium'') 
importance in terms of their contribution to extinction risk of all 
coral species across their ranges. These qualitative rankings 
considered: (1) The severity of the threat; (2) the geographic scope of 
the threat; (3) the level of certainty that corals in general (given 
the paucity of species-level information) are affected by each threat; 
(4) the projections of potential changes in the threat; and (5) the 
impacts of the threat on each species. Global climate change directly 
influences two of the three highest ranked threats, ocean warming and 
ocean acidification, and indirectly (through ocean warming) influences 
the remaining highest ranked threat, disease.
    We identified nine threats (see Table 1) as posing either current 
or future extinction risk to the proposed corals. However, the SRR 
identified 19 threats that affect corals. The ten threats not included 
in Table 1 did not rank highly in their contribution to extinction 
risk, although they do adversely affect the species. Ocean warming, 
ocean acidification, and disease are overarching threats of high or 
medium-high importance when evaluating the extinction risk of the 
proposed species. These impacts are currently occurring, and are 
expected to worsen, posing increasingly severe effects on the species 
considered in this final rule. Other threats are of medium or medium-
low importance when evaluating extinction risk because their effects 
are largely indirect and/or local to regional in spatial scale. These 
include trophic effects of fishing, sea-level rise, and water quality 
issues related to sedimentation and nutrients. The remaining threats 
can be locally acute, but because they affect limited geographic areas, 
they are of low importance when evaluating extinction risk. Examples in 
this category are predation or collection for the ornamental trade 
industry. These threats are more significant to certain species, such 
as those with naturally low abundance and/or those at severely depleted 
population levels. However, none of the species in this final rule can 
be characterized as such.
    Table 1. The nine most important threats contributing to extinction 
risk for corals in general and ordered according to importance. The 
threat is paired with its corresponding ESA section 4 factor in the 
last column.

[[Page 53886]]



----------------------------------------------------------------------------------------------------------------
                 Threat                               Importance                       Section 4 factor
----------------------------------------------------------------------------------------------------------------
Ocean Warming..........................  High...............................  E.
Disease................................  High...............................  C.
Ocean Acidification....................  Medium-High........................  E.
Trophic Effects of Fishing.............  Medium.............................  A.
Sedimentation..........................  Low-Medium.........................  A and E.
Nutrients..............................  Low-Medium.........................  A and E.
Sea-Level Rise.........................  Low-Medium.........................  A.
Predation..............................  Low................................  C.
Collection and Trade...................  Low................................  B.
----------------------------------------------------------------------------------------------------------------

    Some comments (e.g., Comment 26) suggested that local threats, such 
as sedimentation, are more important locally to species' extinction 
risk than the higher rated threats. In the proposed rule, we 
acknowledged that some of the local threats have been the cause of mass 
coral mortality in particular locations. Further, supplemental 
information provides evidence that local threats, such as overfishing 
and disease, have actually been more significant drivers of past coral 
reef species decline, particularly in the Caribbean (Jackson et al., 
2014). However, we must evaluate all threats that pose an extinction 
risk to the proposed species over the foreseeable future. Given the 
predicted impacts of climate-related threats over the foreseeable 
future, we maintain the relative importance ranking of the threats to 
reef-building corals generally. However, we acknowledge that lower 
importance threats also pose significant risk to individual species in 
certain locations.

Foreseeable Future

    In the proposed rule, we established that the appropriate period of 
time corresponding to the foreseeable future is a function of the 
particular types of threats, the life-history characteristics, and the 
specific habitat requirements for the coral species under 
consideration. The timeframe corresponding to the foreseeable future 
takes into account the time necessary to provide for the conservation 
and recovery of each threatened species (e.g., recruitment rate, growth 
rate, etc.) and the ecosystems upon which they depend, but is also a 
function of the reliability of available data regarding the identified 
threats and extends only as far as the data allow for making reasonable 
predictions about the species' response to those threats. As is 
discussed further in the Foreseeable Future and Current and Future 
Environmental Conditions subsections of the Risk Analysis section 
below, the period of time over which individual threats and responses 
may be projected varies according to the nature of the threat and the 
type of information available about that threat and the species' likely 
response. As described below, the more vulnerable a coral species is to 
the high importance threats (i.e., ocean warming, diseases, ocean 
acidification), the more likely the species is at risk of extinction, 
either now or within the foreseeable future. The threats related to 
global climate change (e.g., bleaching from ocean warming, ocean 
acidification) pose the greatest potential extinction risk to corals 
and have been evaluated with sufficient certainty out to the year 2100.
    Comment 38 provides a summary of the comments we received on the 
determination of foreseeable future in the proposed rule and supporting 
documents as extending out to the year 2100. Many comments criticized 
the use of 2100 because they considered it to be too far into the 
future. We do not agree that 2100 is too far in the future to be 
considered foreseeable as it pertains to projections regarding climate-
change related threats. As described in detail in the Global Climate 
Change--General Overview section, the IPCC Fifth Assessment Report 
(AR5), Climate Change 2013: The Physical Science Basis (IPCC, 2013), 
commonly referred to as the Working Group I Report (WGI), is a 
continuation of AR4. Most of AR5 WGI's models also use 2100 as the end-
point (some models go beyond 2100) and the supplemental information 
included in AR5 reinforces our original basis for defining the 
foreseeable future as the period of time from the present to the year 
2100 (IPCC, 2013). That is, the foreseeable future is not defined as 
the year 2100, but rather as the time period from the present to the 
year 2100, with increasing uncertainty in climate change projections 
over that time period. So while precise conditions during the year 2100 
are not reasonably foreseeable, the general trend in conditions during 
the period of time from now to 2100 including the period 2081 to 2100 
is reasonably foreseeable as a whole, although less so through time. 
Because the time period of the present to the year 2100 is strongly 
supported as a reasonably foreseeable timeframe in the climate science 
projections in AR5's WGI, and because the climate-related impacts to 
coral reefs may be substantial within that timeframe, our conclusion 
that 2100 is the appropriate timeframe for purposes of analyzing 
climate change-related threats remains unchanged.

Nine Most Important Threats to Reef-Building Corals

    As described above and shown in Table 1, we considered nine threats 
to be the most important to the current or expected future extinction 
risk of reef-building corals: Ocean warming, disease, ocean 
acidification, trophic effects of reef fishing, sedimentation, 
nutrients, sea-level rise, predation, and collection and trade. 
Vulnerability of a coral species to a threat is a function of 
susceptibility and exposure, considered at the appropriate spatial and 
temporal scales. In this finding, the spatial scale is the current 
range of the species, and the temporal scale is from now through the 
foreseeable future. Susceptibility refers to the response of coral 
colonies to the adverse conditions produced by the threat. 
Susceptibility of a coral species to a threat is primarily a function 
of biological processes and characteristics, and can vary greatly 
between and within taxa. Susceptibility depends on direct effects of 
the threat on the species, and it also depends on the cumulative (i.e., 
additive) and interactive (i.e., synergistic or antagonistic) effects 
of multiple threats acting simultaneously on the species. Exposure 
refers to the degree to which the species is likely to be subjected to 
the threats throughout its range, so the overall vulnerability of a 
coral species to threats depends on the proportion of colonies that are 
exposed to the threats. Thus, the exposure of a species to threats, on 
a range-wide scale, is a function of physical processes and 
characteristics that affect the frequency or degree to which individual 
colonies experience the threats and the ability of its spatial and 
demographic traits to affect its overall vulnerability. A species may 
not necessarily be highly vulnerable to a threat even when it is highly 
susceptible to the threat, if exposure is low over the appropriate

[[Page 53887]]

spatial and temporal scales. Consideration of the appropriate spatial 
and temporal scales is particularly important, because of potential 
high variability in some threats over the large spatial scales. The 
nine most important threats are summarized below, including general 
descriptions of susceptibility and exposure. Species-specific threat 
susceptibilities are described in the Species-specific Information and 
Determinations section.

Global Climate Change--General Overview

    Several of the most important threats contributing to the 
extinction risk of corals are related to global climate change. The 
main concerns regarding impacts of global climate change on coral reefs 
generally, and on the proposed corals in particular, are the magnitude 
and the rapid pace of change in GHG concentrations (e.g., carbon 
dioxide (CO2) and methane) and atmospheric warming since the 
Industrial Revolution in the mid-19th century. These changes are 
increasing the warming of the global climate system and altering the 
carbonate chemistry of the ocean (ocean acidification), which affects a 
number of biological processes in corals, including secretion of their 
skeletons. The description and analysis of global climate change in the 
proposed rule and supporting documents were based largely on the IPCC 
AR4, The Physical Science Basis (IPCC, 2007) and supporting literature. 
Supplemental information gathered during the public engagement period 
shows that global temperatures continue to increase and that 
temperature patterns differ regionally.
    As summarized in Comment 11, we received many comments on our 
analysis of global climate change in the proposed rule. Some commenters 
asserted that we did not adequately portray the level of uncertainty 
associated with the available climate change models. Others provided 
information that global GHG emissions and global temperatures continue 
to rise unabated. Additionally, significant supplemental information 
has become available on global climate change since the proposed rule, 
specifically, AR5's WGI (IPCC, 2013), and its companion report, Climate 
Change 2014: Impacts, Adaptation, and Vulnerability, commonly referred 
to as the Working Group II Report (WGII; IPCC, 2014).
    The IPCC has summarized the major sources of uncertainty associated 
with AR5's WGI projections of global climate change as: (1) The 
projected rate of increase for GHG concentrations; (2) strength of the 
climate's response to GHG concentrations; and (3) large natural 
variations. The warming rate slow-down (or ``hiatus'' discussed in the 
Threats Evaluation--Ocean Warming section) since 1998 is an example of 
a large natural variation that was not predicted by the models at that 
time. However, AR4's projections are built on scientifically sound 
principles, and they fairly simulate many large-scale aspects of 
present-day conditions, and thereby provided the best available 
information on climate change at the time the proposed rule was 
published. Overall uncertainty is not necessarily any greater in AR5 
than in AR4, but rather the uncertainty is understood better and 
expressed more clearly in AR5's WGI (IPCC, 2007; IPCC, 2013; Knutti and 
Jan Sedl[aacute]cek, 2012). AR5's WGI represents the largest synthesis 
of global climate change physical science ever compiled, and a 
substantial advance from AR4. WGI is divided into four sections that 
examine observations, drivers, understanding, and projections of 
changes to the global climate system. The primary results of these four 
sections relevant to this rule are summarized below; then a summary of 
the potential impacts to corals resulting from the IPCC climate change 
scenario that we consider to be the most impactful to corals is 
provided in the RCP8.5 Projections section below, with a focus on ocean 
warming and acidification, two of the most important threats to corals.
    The first section of WGI considers observations of changes in the 
climate system, which refers to description of past climate patterns, 
and the certainty associated with the same. The overall conclusion of 
this section is that warming of the climate system is unequivocal and 
since the 1950s, many of the observed changes are unprecedented over 
decades to millennia. With regard to ocean warming, it is ``virtually 
certain'' that the upper ocean (0-700 m) warmed from 1971 to 2010. With 
regard to ocean acidification, it is ``very likely'' that the pH of 
surface ocean waters has decreased as a result of ocean uptake of 
anthropogenic CO2 from the atmosphere. With regard to sea-
level rise, it is ``virtually certain'' that the global mean sea level 
rose by 19 cm from 1901 to 2010 (IPCC, 2013).
    The second section of WGI considers drivers of changes in the 
climate system, which refers to explanations of factors forcing climate 
patterns. Natural and anthropogenic substances and processes that alter 
the Earth's energy budget are drivers of climate change. In AR5, 
radiative forcing (RF, measured in watts per square meter, W/m\2\) 
quantifies energy fluxes caused by changes in these drivers relative to 
the year 1750. Increasing RF leads to surface warming, and decreasing 
RF leads to surface cooling. The concentration of CO2 in the 
atmosphere is the dominant anthropogenic driver. Higher atmospheric 
CO2 results in: Ocean warming via the greenhouse effect, 
ocean acidification via oceanic uptake of CO2, and rising 
sea levels via ice melting and thermal expansion. Patterns in solar 
activity and major volcanic eruptions are the two dominant natural 
drivers. Solar activity can either increase or decrease RF, whereas 
major volcanic eruptions only decrease RF. Current total RF relative to 
1750 is positive, and has led to an uptake of energy by the climate 
system. The largest contribution to current total RF is the increasing 
atmospheric concentration of CO2 since 1750, most of which 
has been anthropogenic CO2 emitted since 1860, and the mean 
rate of increase in CO2 is unprecedented in the past 20,000 
years. Current CO2 levels (~400 ppm) will result in 
continued warming even if anthropogenic emissions went to zero now 
(this is referred to as ``commitment'' to future warming from the 
CO2 build-up already in the atmosphere), but reducing 
emissions now would strongly influence the levels of future warming 
(IPCC, 2013).
    The third section of WGI describes past climate patterns to 
understand the changes in the climate system. It is ``extremely 
likely'' that human activities caused more than half of the observed 
increase in global average surface temperature from 1951 to 2010. 
Anthropogenic GHGs have ``very likely'' made a substantial contribution 
to upper-ocean warming (above 700 m) observed since the 1970s. It is 
also ``very likely'' that oceanic uptake of anthropogenic 
CO2 has reduced surface water pH. The anthropogenic ocean 
warming observed since the 1970s has contributed to global sea-level 
rise over this period through ice melting and thermal expansion (IPCC, 
2013).
    The fourth section of WGI uses projected changes in the climate 
system to model potential patterns of future climate. WGI uses a new 
set of four representative concentration pathways (RCP) that provide a 
standard framework for consistently modeling future climate change. 
These replace the old Special Report on Emissions Scenarios (SRES) 
system used in prior assessments. The new RCPs are named according to 
increases in radiative forcing (RF) relative to the 1986-2005 average 
by the year 2100 of 2.6, 4.5, 6.0,

[[Page 53888]]

and 8.5 W/m\2\, RCP2.6, RCP4.5, RCP6.0, and RCP8.5. The four new 
pathways have atmospheric CO2 equivalents of 421 (RCP2.6), 
538 (RCP4.5), 670 (RCP6.0), and 936 ppm (RCP 8.5) in 2100, and follow 
very different trajectories to reach those endpoints. The purpose of 
the RCPs was to explicitly explore the impact of different climate 
policies in addition to the no-climate-policy scenarios explored in the 
earlier scenarios (Van Vuuren et al., 2011). The four new pathways were 
developed with the intent of providing a wide range of total climate 
forcing to guide policy discussions and specifically include one 
mitigation pathway leading to a very low forcing level (RCP2.6), two 
stabilization pathways (RCP4.5 and RCP6), and one pathway with 
continued high GHG emissions (RCP8.5).
    The RCP method more strongly represents the physical processes 
underlying climate change, and various factors affecting GHG emissions 
globally, than previous methods. WGI adjusts the likely global surface 
warming that would result from a doubling of atmospheric CO2 
to 1.5-4.5 [deg]C (compared to AR4's estimate of 2.0-4.5 [deg]C), due 
to improved understanding of the climate system, the extended 
temperature record in the atmosphere and ocean, and new estimates of 
radiative forcing to GHG concentrations. Taken together, the four new 
pathways project wide ranges of increases in ocean warming, ocean 
acidification, and sea level rise globally throughout the 21st century 
with conditions seen in RCP 2.6-6.0 requiring significant changes in 
anthropogenic GHG emissions (IPCC, 2013).
    The proposed rule and supporting documents assumed that AR4's 
highest-emission scenario A1FI was the most likely to occur for two 
reasons: (1) Recent annual GHG emission growth rates had exceeded the 
GHG emission growth rates in A1F1 (except 2009 when the global 
recession slowed growth); and (2) there were no indications that major 
reductions in GHG emissions would occur in the near to mid-term future 
(decades) through national or international policies or major changes 
in the global fossil fuel economy (Brainard et al., 2011). Recent 
annual GHG emission growth rates (except 2009) exceed the GHG emission 
growth rates in RCP8.5 (Le Qu[eacute]r[eacute] et al., 2013). While the 
President's Climate Action Plan and intensified international climate 
negotiations may change global emissions trajectories, we make the 
conservative assumption to evaluate RCP8.5, and its projections for 
ocean warming and ocean acidification, in our assessment of extinction 
risk for the corals in the final rule. RCP8.5 is the scenario with the 
highest GHG emissions rate and subsequent future GHG levels; thus it 
would be the most impactful to corals through ocean warming and ocean 
acidification. However, should another of the IPCC RCPs ultimately be 
realized, the negative impacts to corals would be lower.
    As described above, we received and collected significant 
supplemental information regarding our consideration of global climate 
change in the proposed rule. Additional observations, data, and testing 
have produced better models and a greater understanding of the 
uncertainty inherent in climate change projections. Annual GHG emission 
rates continue to climb to record levels, and the last decade has been 
the warmest on record, underscoring the proposed rule's conclusions 
about climate change threats to reef-building corals. We conclude that 
the supplemental information supports the central premise of the 
proposed rule that global climate change-related threats have already 
caused widespread impacts to corals and coral reefs and these impacts 
will become increasingly severe from now to 2100, with correspondingly 
severe consequences for corals and coral reefs. However, we acknowledge 
that the interpretation of future climate change threats to corals and 
coral reefs is associated with complexity and uncertainty, and that 
precise effects on individual species of reef-building corals are 
difficult to determine. Species-specific threat susceptibilities of 
each of the 65 species in this final rule to the threats resulting from 
global climate change are described in the Species-specific Information 
and Determinations section below.

RCP8.5 Projections

    Because we have determined that RCP8.5 is the most impactful 
pathway to corals, we provide a summary of RCP8.5's projections over 
the foreseeable future for ocean warming and ocean acidification (IPCC, 
2013). Where possible, projections are provided for the near-term (to 
mid-century) and long-term (to 2100), and globally and regionally 
(Indo-Pacific and Caribbean). Implications for coral reefs are also 
described.
    Ocean Warming. Under RCP8.5, annual averaged, globally averaged, 
surface ocean temperature is projected to increase by approximately 0.7 
[deg]C by 2030 and 1.4 [deg]C by 2060 compared to the 1986-2005 
average, with the 10 to 90 percent range increasing over that time 
period to approximately +/-0.7 [deg]C by 2060 (IPCC, 2013; WGI Figure 
11.19). Projected changes in annual mean ocean temperature between 60 
[deg]N and 60 [deg]S latitude in 2081-2100 are shown in WGI Figure 
12.12. Under RCP8.5, annual mean surface ocean temperature between 60 
[deg]N and 60 [deg]S latitude is projected to increase by approximately 
3.5 [deg]C by 2081-2100 compared to the 1986-2005 average (IPCC, 2013; 
WGI Figure 12.12). A different graph using the same data shows global 
annual mean surface ocean temperature is projected to increase by 
approximately 3.5 [deg]C by 2081-2100 compared to the 1986-2005 
average, with 5 to 95 percent range of +/-1-1.5 [deg]C (IPCC, 2013; 
Figure AI.SM8.5.4). Thus, RCP8.5 projects that global annual mean ocean 
surface temperatures will increase by approximately 0.4-1 [deg]C by 
2030, approximately 0.7-2 [deg]C by 2060, and approximately 2-5 [deg]C 
by 2081-2100 (IPCC, 2013).
    Projected changes in Indo-Pacific annual median ocean surface 
temperatures (i.e., WGI's West Indian Ocean, North Indian Ocean, 
Southeast Asia, North Australia, and Pacific Islands regions), and 
Caribbean annual median land and ocean combined surface temperatures, 
compared to the 1986-2005 average are shown in the figures in WGI's 
Annex I's Supplementary Material for RCP8.5 for these six WGI regions, 
which together cover the ranges of the species included in this final 
rule. The figures include graphs in the upper right showing the 
projected median temperature increase to 2100 under RCP8.5, the 25 to 
75 percent range, and the 5 to 95 percent range. The figures also 
includes maps of each region showing projected changes spatially under 
RCP8.5 for the time periods 2016-2035, 2046-2065, and 2081-2100, and 
for the 25 percent, 50 percent, and 75 percent projections under RCP8.5 
for each of these time periods. For the Caribbean, the range of 
projections spanned by the 25, 50, and 75 percent range maps are: For 
2016-2035, increases of 0.5-1.0 [deg]C; for 2046-2065, increases of 
1.0-3.0 [deg]C; and for 2081-2100, increases of 2.0-4.0 [deg]C. Spatial 
variability in the projections consists mostly of larger increases in 
the Greater Antilles and Jamaica, and lower increases in the Lesser 
Antilles and the Bahamas (Figure AI.SM8.5.44). The percent ranges in 
the projections described above are from the maps and are for the 25 to 
75 percent range, however range of projections within the 5 to 95 
percent range are considerably greater, as shown in the bar-and-whisker

[[Page 53889]]

graph in the upper right of each figure (IPCC, 2013).
    For the Indo-Pacific (WGI's West Indian Ocean, North Indian Ocean, 
Southeast Asia, North Australia, and Pacific Islands regions), the 
range of projections spanned by the 25, 50, and 75 percent range maps 
are: For 2016-2035, increases of 0.0-1.0 [deg]C; for 2046-2065, 
increases of 1.0-3.0 [deg]C; and for 2081-2100, increases of 2.0-5.0 
[deg]C. Spatial variability in the projections consists mostly of 
larger increases in the Red Sea, Persian Gulf, and the Coral Triangle, 
and lower increases in the central and eastern Indian Ocean and south-
central Pacific (Figures AI.SM8.5.92, 116, 124, 132, and 140). The 
percent ranges in the projections described above are from the maps and 
are for the 25 to 75 percent range, however range of projections within 
the 5 to 95 percent range are considerably greater, as shown in the 
bar-and-whisker graph in the upper right of each figure (IPCC, 2013).
    To summarize ocean warming projections, RCP8.5 projects annual 
median ocean surface temperature increases for the Indo-Pacific, and 
annual median land and ocean combined surface temperature increases for 
the Caribbean. Projected median temperatures, and associated 25 to 75 
percent range and 5 to 95 percent range, are provided for the time 
periods of 2016-2035, 2046-2065, and 2081-2100. We interpret these 
projections as follows: (1) Global annual median ocean surface 
temperatures are likely to rise approximately 2-5 [deg]C by 2081-2100, 
exacerbating the impacts of ocean warming on reef-building corals; (2) 
these global mean projections are not necessarily representative of 
ocean surface temperature conditions throughout the ranges and habitats 
of the species in this final rule through the foreseeable future, due 
to spatial variability and statistical range of the RCP8.5 ocean 
warming projections described above for the Indo-Pacific and Caribbean 
regions; and (3) ocean surface temperature conditions in the 
foreseeable future within the ranges of the species in this final rule 
are assumed to vary spatially at the coarse spatial scales shown in WGI 
for the Indo-Pacific and Caribbean regions, and more so at finer 
spatial scales, and to fall within the statistical ranges projected for 
the Indo-Pacific and Caribbean regions.
    Ocean Acidification. Under RCP8.5, mean surface pH in the tropics 
(20 [deg]N to 20 [deg]S) is projected to decline from the current pH of 
approximately 8.05 to approximately 7.95 by 2050, and to approximately 
7.75 by 2100, or a reduction of 0.31 (statistical range of 0.30 to 
0.32) by 2100 (IPCC, 2013; WGI Figure 6.28a). Projected changes in 
global surface pH in the 2090s compared to the 1990s under RCP8.5 are 
shown in the map in WGI Figure 6.28b. In the tropical Indo-Pacific, 
decreases of 0.25 to 0.40 are projected, with the lower decreases in 
the central and eastern Pacific, and the higher decreases in the GBR 
area and the northern Philippines, while most of the Caribbean is 
projected to decrease in pH by 0.30 to 0.35. The pH reductions 
associated with RCP8.5 are projected to result in declining aragonite 
saturation states, as shown in WGI Figure 6.29. Projected median 
surface aragonite saturation states of the world's oceans are shown for 
2050 and 2100 in Figure 6.29d and f respectively, and by depth for the 
Atlantic and Pacific Oceans in 2100 in Figure 6.29c and e respectively. 
Surface aragonite saturation states in the tropical Indo-Pacific and 
Caribbean are projected to decline from current levels of over 3, to 
less than 2.5 by 2100, with similar spatial patterns as for pH 
reductions (IPCC, 2013; WGI Figure 6.29). Statistical range is not 
provided for aragonite saturation state, but we assume it to be similar 
to that associated with pH projections. As shown in Figures 6.28 and 
6.29, spatial variability is projected under RCP8.5 for both pH and 
aragonite saturation state reductions over the foreseeable future 
within the ranges of the species included in this final rule (IPCC, 
2013).
    We interpret RCP8.5's ocean acidification projections as follows: 
(1) Mean surface pH in the tropics is projected to decline by 
approximately 0.31 to approximately 7.75 by 2100, with a subsequent 
large decline in aragonite saturation state in surface tropical waters, 
exacerbating the impacts of ocean acidification on reef-building 
corals; (2) surface pH and aragonite saturation state conditions 
throughout the ranges of the species in this final rule through the 
foreseeable future are not necessarily represented by these mean 
projections, due to the spatial variability within the Indo-Pacific and 
Caribbean regions, and the statistical range of the RCP8.5 ocean 
acidification projections; and (3) surface pH and aragonite saturation 
state conditions in the foreseeable future within the ranges of the 
species in this final rule are assumed to vary spatially at the coarse 
spatial scales shown in WGI for the Indo-Pacific and Caribbean regions, 
and more so at finer spatial scales, and to fall within the statistical 
ranges projected for the Indo-Pacific and Caribbean regions.
    Implications for Coral Reef Ecosystems. AR5's WGII Report describes 
the effects of WGI's climate change projections on the world's 
ecosystems, including coral reefs. The report includes a description of 
``Projected Impacts'' on coral reefs of all four WGI pathways combined, 
and a general overview of projected impacts to coral reefs. While this 
information does not specifically describe projected impacts of RCP8.5 
to coral reefs by 2100, it strongly suggests that the projected impacts 
of ocean warming and ocean acidification will increase (IPCC, 2014). 
Likewise, the recent U.S. National Climate Assessment (NCA) report 
describes the effects of projected climate change on United States 
ecosystems, including coral reefs. Chapter 24 of the report includes a 
brief and general description of projected climate change without 
specifically examining any particular pathway (Doney et al., 2014). As 
with WGII, while the NCA report does not specifically describe 
projected impacts of RCP8.5 to coral reefs by 2100, it strongly 
suggests that the projected impacts of ocean warming and ocean 
acidification will increase on United States coral reefs.
    Recent papers specifically address future changes in Indo-Pacific 
and Caribbean coral reef ecosystems resulting from RCP8.5's projections 
of combined ocean warming and ocean acidification, including Couce et 
al. (2013a) and van Hooidonk et al. (2014). Couce et al. (2013a) uses 
RCP8.5's ocean warming and ocean acidification projections to develop 
predictions of ``average change in suitability'' of coral reef habitat 
by 2070, concluding that declines in conditions will be driven 
primarily by ocean warming, and vary spatially within the ranges of the 
species included in this final rule. Couce et al. (2013) predicts 
marked declines in environmental suitability for shallow coral reef 
habitats across the equatorial western Pacific and adjacent areas 
(e.g., Coral Triangle) by 2070, and generally less favorable conditions 
elsewhere on Indo-Pacific and Caribbean coral reefs. Some coral reef 
areas show little or no change in environmental suitability by 2070, 
including portions of the western Indian and central Pacific Oceans, 
likely because seawater temperatures are moderated by physical factors 
such as higher latitudes or upwelling but aragonite saturation states 
are suitable (Couce et al., 2013a; Fig. 1e). Many species included in 
this final rule occur in areas of the western Indian and central 
Pacific Oceans predicted to have

[[Page 53890]]

little or no change in environmental suitability by 2070. Notably, the 
paper concluded the detrimental effect of higher ocean warming appears 
to strongly outweigh the impacts of lower aragonite saturation states 
for tropical shallow water coral reefs (Couce et al., 2013a).
    van Hooidonk et al. (2014) also applies RCP8.5's ocean warming and 
ocean acidification projections to predict ``when severe coral 
bleaching events start to occur annually, and of changes in aragonite 
saturation state'' over the 21st century. The paper concludes that 90 
percent of all coral reefs are projected to experience severe bleaching 
annually by 2055, that five percent declines in calcification are 
projected for all reef locations by 2034, with the predicted changes in 
conditions varing spatially across the geographic ranges of the species 
included in this final rule. These authors predicted that the most 
rapid increases in ocean warming will occur in the western equatorial 
Pacific, the slowest in the Indian Ocean, eastern Pacific Ocean, and 
high latitude areas, and intermediate elsewhere (van Hooidonk et al., 
2014; Fig 1a). The most rapid declines in aragonite saturation state 
are predicted for the same general areas as the slowest warming, the 
slowest declines in aragonite saturation state in roughly the same 
areas as the most rapid warming, and intermediate elsewhere in the 
Indo-Pacific and in the Caribbean (van Hooidonk et al., 2014; Fig 1d). 
One of the paper's conclusions is that there are no real refugia for 
coral reefs to the combined threats of higher ocean warming and lower 
aragonite saturation states (van Hooidonk et al., 2014).
    Several points to consider when interpreting Couce et al. (2013a) 
and van Hooidonk et al. (2014) are: (1) The different results and 
conclusions are likely due to the different methods, and illustrate the 
sensitivity and variability in predicting the impacts of projected 
changes in climate on coral reefs; (2) both papers used very coarse 
spatial scales (1[deg] x 1[deg] cells, or >10,000 km\2\ at the 
Equator), thus each cell can include many different reefs that 
collectively represent diverse coral communities and habitats, which in 
turn can affect the local spatial and temporal patterns of coral 
responses to ocean warming and acidification; (3) both papers predict 
high spatial variability in future conditions across coral reefs, and 
both show the western equatorial Pacific as having the most degraded 
future conditions, and parts of the Indian Ocean, central Pacific, and 
some outlying areas as having less degraded future conditions; and (4) 
neither paper analyzed the impacts of future climate change on 
individual coral species.
    In conclusion, RCP8.5 projects impacts to global coral reef 
ecosystems over the foreseeable future from the combined effects of 
increased ocean temperature and ocean acidification, the effects of 
which are likley to be compounded by increasing coral disease, trophic 
effects of fishing, land-based sources of pollution, and other threats 
to corals. However, projecting species-specific responses to global 
threats is complicated by several physical and biological factors: (1) 
Global projections of changes to ocean temperatures and acidification 
over the foreseeable future are associated with three major sources of 
uncertainty (GHG emissions assumptions, strength of the climate's 
response to GHG concentrations, and large natural variations); (2) 
there is spatial variability in projected environmental conditions 
across the ranges of the species in this final rule at any given point 
in time; and (3) species-specific responses depend on many biological 
characteristics, including, at a minimum, distribution, abundance, life 
history, susceptibility to threats, and capacity for acclimatization. 
The available species-specific information on how species in this final 
rule respond to climage change is limited. Therefore, analysis of the 
biological characteristics on a case-by-case basis is emphasized in 
considering a species' vulnerability to extinction.

Ocean Warming (High Importance Threat, ESA Factor E)

    Ocean warming is considered under ESA Factor E--other natural or 
manmade factors affecting the continued existence of the species--
because the effect of the threat results from human activity and 
affects individuals of the species directly, and not their habitats. In 
the proposed rule, we described the threat from ocean warming as 
follows. Mean seawater temperatures in reef-building coral habitat in 
both the Caribbean and Indo-Pacific have increased during the past few 
decades, and are predicted to continue to rise between now and 2100. As 
also described in the proposed rule, the frequency of warm-season 
temperature extremes (warming events) in reef-building coral habitat in 
both the Caribbean and Indo-Pacific has increased during the past two 
decades, and it is also predicted to increase between now and 2100.
    Ocean warming is one of the most important threats posing 
extinction risks to the proposed coral species; however, individual 
susceptibility varies among species. The primary observable coral 
response to ocean warming is bleaching of adult coral colonies, wherein 
corals expel their symbiotic zooxanthellae in response to stress. For 
many corals, an episodic increase of only 1 [deg]C-2 [deg]C above the 
normal local seasonal maximum ocean temperature can induce bleaching. 
Corals can withstand mild to moderate bleaching; however, severe, 
repeated, or prolonged bleaching can lead to colony death. Coral 
bleaching patterns are complex, with several species exhibiting 
seasonal cycles in symbiotic dinoflagellate density. Thermal stress has 
led to bleaching and associated mass mortality in many coral species 
during the past 25 years. In addition to coral bleaching, other effects 
of ocean warming detrimentally affect virtually every life-history 
stage in reef-building corals. Impaired fertilization, developmental 
abnormalities, mortality, impaired settlement success, and impaired 
calcification of early life phases have all been documented. In the 
proposed rule, we relied heavily on AR4 in evaluating extinction risk 
from ocean warming because it contained the most thoroughly documented 
and reviewed assessments of future climate and represented the best 
available scientific information on potential future changes in the 
earth's climate system. Emission rates in recent years have met or 
exceeded levels predicted by AR4's worst-case scenarios, resulting in 
all scenarios underestimating the projected climate condition.
    Exposure of colonies of a species to ocean warming can vary greatly 
across its range, depending on colony location (e.g., latitude, depth, 
bathymetry, habitat type, etc.) and physical processes that affect 
seawater temperature and its effects on coral colonies (e.g., winds, 
currents, upwelling shading, tides, etc.). Colony location can moderate 
exposure of colonies of the species to ocean warming by latitude or 
depth, because colonies in higher latitudes and/or deeper areas are 
usually less affected by warming events. Deeper areas are generally 
less affected typically because lower irradiance reduces the likelihood 
of warming-induced bleaching. Also, some locations are blocked from 
warm currents by bathymetric features, and some habitat types reduce 
the effects of warm water, such as highly fluctuating environments. 
Physical processes can moderate exposure of colonies of the species to 
ocean warming in many ways, including processes that increase mixing 
(e.g., wind, currents, tides), reduce seawater temperature (e.g.,

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upwelling, runoff), or increase shading (e.g. turbidity, cloud cover). 
Exposure of colonies of a species to ocean warming will likely vary 
annually and decadally, while increasing over time, because: (1) 
Numerous annual and decadal processes that affect seawater temperatures 
will continue to occur in the future (e.g., inter-decadal variability 
in seawater temperatures and upwelling related to El-Ni[ntilde]o 
Southern Oscillation); and (2) ocean warming is predicted to 
substantially increase by 2100.
    Multiple threats stress corals simultaneously or sequentially, 
whether the effects are cumulative (the sum of individual stresses) or 
interactive (e.g., synergistic or antagonistic). Ocean warming is 
likely to interact with many other threats, especially considering the 
long-term consequences of repeated thermal stress, and that ocean 
warming is expected to continue to increase over the foreseeable 
future. Increased seawater temperature can lower resistance to coral 
diseases and reduce coral health and survivorship. Coral disease 
outbreaks often have either accompanied or immediately followed 
bleaching events, and also follow seasonal patterns of high seawater 
temperatures. The effects of greater ocean warming (e.g., increased 
bleaching, which kills or weakens colonies) are expected to interact 
with the effects of higher storm intensity (e.g., increased breakage of 
dead or weakened colonies), resulting in an increased rate of coral 
declines. Likewise, ocean acidification and nutrients may reduce 
thermal thresholds to bleaching, increase mortality, and slow recovery.
    There is also mounting evidence that warming ocean temperatures can 
have direct impacts on early life stages of corals, including abnormal 
embryonic development at 32 [deg]C and complete fertilization failure 
at 34 [deg]C for one Indo-Pacific Acropora species. In addition to 
abnormal embryonic development, symbiosis establishment, larval 
survivorship, and settlement success have been shown to be impaired in 
Caribbean brooding and broadcasting coral species at temperatures as 
low as 30 [deg]C-32 [deg]C. Further, the rate of larval development for 
spawning species is appreciably accelerated at warmer temperatures, 
which suggests that total dispersal distances could also be reduced, 
potentially decreasing the likelihood of successful settlement and the 
replenishment of extirpated areas.
    Finally, warming will continue causing increased stratification of 
the upper ocean because water density decreases with increasing 
temperature. Increased stratification results in decreased vertical 
mixing of both heat and nutrients, leaving surface waters warmer and 
nutrient-poor. While the implications for corals and coral reefs of 
these increases in warming-induced stratification have not been well 
studied, it is likely that these changes will both exacerbate the 
temperature effects described above (e.g., increase bleaching and 
decrease recovery) and decrease the overall net productivity of coral 
reef ecosystems (e.g., fewer nutrients) throughout the tropics and 
subtropics.
    Overall, there is ample evidence that climate change (including 
that which is already committed to occur from past GHG emissions and 
that which is reasonably certain to result from continuing and future 
emissions) will follow a trajectory that will have a major impact on 
corals. There has been a recent research emphasis on the processes of 
acclimatization and adaptation in corals, but in the proposed rule we 
determined that, taken together, the body of research was inconclusive 
as to how these processes may affect individual corals' extinction 
risk, given the projected intensity and rate of ocean warming. As 
detailed in Comments 12-16, we received numerous comments related to 
ocean warming threats to corals that focused on the following aspects: 
(1) General future projections of ocean warming levels; (2) accounting 
for spatial variability; (3) the future decline of coral reefs because 
of increasing GHG emissions; (4) the possibility of wide ranging 
responses by coral reef ecosystems; (5) the specific effects of ocean 
warming on reef-building corals; and (6) the capacity of reef-building 
corals for acclimatization and adaptation to ocean warming.
    With regard to the future projections of global climate change, the 
proposed rule and supporting documents assumed that AR4's highest-
emission scenario A1FI was the most likely. As discussed in Global 
Climate Change--General Overview, we assume that for corals RCP8.5 is 
the most impactful pathway for present to the year 2100. Ocean warming 
projections and implications for coral reefs are described above in the 
RCP8.5 Projections section.
    Comment 12 also criticized our lack of consideration of the post-
1998 hiatus in global warming. The proposed rule did not consider this 
phenomenon as the issue was only emerging during the time the proposed 
rule was drafted. However, because supplemental information has become 
available since that time, we consider it here. Despite unprecedented 
levels of GHG emissions in recent years, a slow-down in global mean 
surface air temperature warming has occurred since 1998, which AR5's 
WGI refers to as a ``hiatus.'' Despite this slowdown in warming, the 
period since 1998 is the warmest recorded and ``Each of the last three 
decades has been successively warmer at the Earth's surface than any 
preceding decade since 1850.''
    The slow-down in global mean surface warming since 1998 is not 
fully explained by AR4 or AR5 WGI's models, but is consistent with the 
substantial decadal and interannual variability seen in the 
instrumental record and may result, in part, from the selection of 
beginning and end dates for such analyses. Possible factors in the 
slow-down may include the following: Heat absorption by the deep ocean 
(Guemas et al., 2013; Levitus et al., 2012) facilitated by stronger 
than normal trade winds (England et al., 2014), volcanic eruptions over 
the last decade (Santer et al., 2014), La Ni[ntilde]a-like decadal 
cooling that produces multi-year periods of slower warming than the 
long-term anthropogenic forced warming trend (Benestad, 2012; 
Easterling and Wehner, 2009; Kosaka and Xie, 2013), inherent 
variability within the climate system that cannot currently be modeled, 
and potentially other factors (IPCC, 2013). As explained above, the 
major sources of uncertainty in climate change projections such as AR4 
or AR5's WGI are: (1) The projected rate of increase for GHG 
concentrations; (2) strength of the climate's response to GHG 
concentrations; and (3) large natural variations. The slow-down in 
warming since 1998 is an example of a large natural variation that 
could not be predicted, at least by the models at that time.
    Comment 12 identified several sources of spatial variability in 
ocean warming and requested our consideration of additional 
information. The proposed rule acknowledged both spatial and temporal 
variability in ocean warming and considered the effect that variability 
would have on the proposed corals. However, we acknowledge that 
supplemental information has since become available, and we consider it 
here. Regional and local variability in ocean warming conditions may 
lead to warming-induced bleaching that is more or less severe 
regionally or locally than globally. A hot spot of ocean warming occurs 
in the equatorial western Pacific where regional warming is higher than 
overall warming in the Indo-Pacific, exposing corals and coral reefs in 
this area to a higher risk of warming-induced bleaching. The hot spot 
overlaps the Coral Triangle (Couce et al., 2013b; Lough, 2012; Teneva 
et al., 2012; van

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Hooidonk et al., 2013b). Several other areas in the Indo-Pacific have 
been identified as having lower than average warming, including the 
western Indian Ocean, Thailand, the southern GBR, central French 
Polynesia, and the eastern equatorial Pacific, potentially resulting in 
relatively lower risk of warming-induced coral bleaching in these areas 
(Couce et al., 2013b; van Hooidonk et al., 2013b). Spatial variability 
in ocean warming is lower in the much smaller Caribbean, and there are 
fewer areas there with lower than average warming (Buddemeier et al., 
2011). The regional and local heterogeneity in ocean warming likely 
results in high variability in coral responses across spatial scales 
(Selig et al., 2010).
    There are several types of temporal variability in ocean warming on 
coral reefs. First, the rate of ocean warming itself changes over time. 
For example, ocean warming has increased in the Indo-Pacific since 
1950, but at different rates at different times (Lough, 2012). Second, 
different periods of ocean warming can result in variability in 
warming-induced bleaching at the same location. For example, a study in 
Thailand showed significant differences in the susceptibility of coral 
taxa to bleaching events between the years 1998 and 2010 and among 
coral species at the same site (Sutthacheep et al., 2013). Spatial 
variability in ocean warming between sites also results in temporal 
variability in ocean warming impacts, as the different areas are 
subsequently affected at different rates into the future (van Hooidonk 
et al., 2013b). For example, a recent study found that Australian 
subtropical reef-building coral communities are affected by ocean 
warming more slowly than tropical reef-building coral communities, 
resulting in slower rates of changes in the subtropical than tropical 
communities (Dalton and Roff, 2013). These studies underscore the 
temporal variability of ocean warming and warming-induced bleaching 
across the ranges of reef-building coral species, complicating the 
interpretation of the effects of ocean warming on any given coral 
species across its range and over time.
    Mesophotic and marginal habitats serving as refugia from ocean 
warming are relatively new and potentially important considerations for 
the vulnerability of coral species to ocean warming. Mesophotic 
habitats continue to be explored, with new surveys finding larger 
habitat areas and greater depth distributions for some reef-building 
corals (Blyth-Skyrme et al., 2013; Bridge and Guinotte, 2012). 
Supplemental information demonstrates the potential for mesophotic 
habitat to provide refugia from ocean warming (Bridge et al., 2013a; 
Smith et al., in press), although it does not always do so (Neal et 
al., 2013). Marginal habitats, such as high latitude sites, upwelling 
regions, and turbid areas like the GBR inner shelf, also may provide 
refugia from ocean warming for some species in some conditions (Browne 
et al., 2012; Couce et al., 2013b; Dalton and Roff, 2013), but not 
others (Lybolt et al., 2011). Taken together, mesophotic and marginal 
habitats may represent a network of refugia from ocean warming for some 
species.
    Comment 14 emphasized both that coral reefs are likely to decline 
sharply in the future because of increasing GHG emissions (e.g., 
Carpenter et al., 2008; Donner, 2009; Frieler et al., 2012; Kiessling 
and Baron-Szabo, 2004) and that a wide range of responses by coral reef 
ecosystems are possible. Studies provided by commenters, and others on 
recent modeling results (Frieler et al., 2012; van Hooidonk and Huber, 
2012; van Hooidonk et al., 2013b) and scientific opinion statements 
(Birkeland et al., 2013; ICRS, 2012) suggest disastrous effects of 
ocean warming, in combination with other threats, on coral reef 
ecosystems. For example, even in AR5 WGI's best-case pathway (RCP2.6) 
where CO2 equivalent concentrations peak at 455 ppm, one 
model suggests that 95 percent of coral reefs will experience annual 
bleaching conditions by the end of the century (van Hooidonk et al. 
2013b). Another model suggests that preserving more than 10 percent of 
coral reefs worldwide would require limiting warming to less than 1.5 
[deg]C above pre-industrial levels. Even assuming high adaptive 
capacity of corals and the more optimistic AR5 pathways, the model 
suggests that one-third of the world's coral reefs are projected to be 
subject to long-term degradation (Frieler et al., 2012). In addition, 
the combined effects of ocean warming and ocean acidification would 
produce even more severe impacts on coral reefs globally (van Hooidonk 
et al., 2013a; Yara et al., 2012).
    These and other studies predict the irreversible disappearance of 
coral reefs on a global scale in the next few decades. However, other 
recent studies suggest that coral reef degradation resulting from 
global climate change threats alone is likely to be a more spatially, 
temporally, and taxonomically heterogeneous process. These studies 
indicate that coral reef ecosystems, rather than disappear entirely as 
a result of future impacts, will likely persist, but with unpredictable 
changes in the composition of coral species and ecological functions 
(Hughes et al., 2012; Pandolfi et al., 2011). Many factors contribute 
to the heterogeneous responses of coral reefs to climate change 
threats, including complexity associated with coral reef habitat, as 
well as the biology of reef-building coral species themselves. As 
described in the Corals and Coral Reefs section, the exceptional 
complexity, extent, and diversity of coral reef habitat increases the 
uncertainty associated with coarse modeling of reef responses to 
climate change threats. Likewise, many aspects of reef-building coral 
biology contribute to complex responses to ocean warming, including 
species-level processes such as capacity for acclimatization and 
adaptation (Palumbi et al., 2014), the potential for range expansion 
(Yamano et al., 2011; Yara et al., 2011), and community-level processes 
such as changes in competition and predation (Cahill et al., 2013; 
Hughes et al., 2012). These different processes occur simultaneously, 
and contribute to highly-variable, complex, and uncertain responses of 
reef-building coral species and in turn coral reefs to climate changes 
threats like ocean warming. Moreover, management of local threats can 
increase resilience of coral reefs to ocean warming and other global 
climate change threats (Jackson et al., 2014; Pandolfi et al., 2011), 
as described further in the Threats Evaluation--Inadequacy of Existing 
Regulatory Mechanisms section.
    Comment 15 focused on the specific effects of ocean warming on 
reef-building corals. The proposed rule described the known specific 
effects of ocean warming as well as the threats that act simultaneously 
or sequentially, and whether the effects are cumulative (the sum of 
individual stresses) or interactive (e.g., synergistic or 
antagonistic). The rapidly growing literature on synergistic effects of 
ocean warming-induced bleaching with other threats demonstrates that 
bleaching is exacerbated by nutrients (Cunning and Baker, 2013; Vega 
Thurber et al., 2013; Wiedenmann et al., 2013), disease is exacerbated 
by warm temperatures and bleaching (Ban et al., 2013; Bruno et al., 
2007; Muller and van Woesik, 2012; Rogers and Muller, 2012), ocean 
warming and acidification may impact corals in opposite but converging 
ways (van Hooidonk et al., 2013a; Yara et al., 2012), and bleaching is 
exacerbated by a variety of physical factors (Yee and Barron, 2010) or 
can be reduced by biological factors (Connolly et al., 2012; Fabricius 
et al., 2013). Other information on species-specifics effects of ocean 
warming is provided in the

[[Page 53893]]

Species-specific Information and Determinations section below.
    Comment 15 focused on the potential capacity of reef-building 
corals for acclimatization and adaptation to ocean warming and provided 
several new studies (Cahill et al., 2013; Guest et al., 2012; Jones and 
Berkelmans, 2010) and some that we considered in the proposed rule 
(Baker et al., 2004; Maynard et al., 2008; Pandolfi et al., 2011). 
Identified mechanisms include symbiont shuffling (Baker, 2012; Cunning 
et al., 2013; Ortiz et al., 2013; Silverstein et al., 2012), symbiont 
shading by host pigments or tissue (Mayfield et al., 2013; Smith et 
al., 2013a), host genotype expression (Baums et al., 2013; Granados-
Cifuentes et al., 2013; Meyer et al., 2011), and host protein 
expression (Barshis et al., 2013; Voolstra et al., 2011). As described 
in the Corals and Coral Reefs section, the dynamic association of host 
coral and symbiotic zooxanthellae and microbes provides potential for 
acclimatization or adaptation of some reef-building coral species to 
environmental changes.
    Many recent studies provide evidence that certain reef-building 
coral communities have acclimated or adapted to ocean warming, at least 
to some degree. The bleaching and mortality of some colonies of a coral 
species on a reef, followed by the recovery of hardier colonies, is the 
process by which acclimatization and adaptation of a species to ocean 
warming occurs. Examples of bleaching, mortality, and recovery provide 
information about the capacity for acclimatization and adaptation. 
Several such examples were provided in the proposed rule and supporting 
documents (Diaz-Pulido et al., 2009; Hueerkamp et al., 2001; Kayanne et 
al., 2002). More recently, many relevant studies have become available 
on the effects of the 1998 bleaching event. For example, in comparisons 
of 1998 and 2010 bleaching events and recovery in southeast Asia, some 
coral species demonstrated more resistance to bleaching in 2010, 
suggesting acclimatization or adaptation to thermal stress (Sutthacheep 
et al., 2013). In a study on an isolated reef in Australia, recovery of 
coral cover occurred within 12 years of the 1998 bleaching event 
(Gilmour et al., 2013). In contrast, studies in the U.S. Virgin Islands 
and Florida demonstrated little if any recovery in the 10 to 12 years 
following the 1998 bleaching event (Rogers and Muller, 2012; Ruzicka et 
al., 2013).
    A recent analysis comparing observed versus predicted coral 
bleaching events suggests that corals may have already responded 
adaptively to some warming since the Industrial Revolution because 
observed bleaching responses are lower than predicted by the warm 
temperature anomalies (Logan et al., 2013). A recent study of fast-
growing, shallow water coral species demonstrated that acclimatization 
and adaptive responses allowed them to inhabit reef areas with water 
temperatures far above their expected tolerances (Palumbi et al., 
2014). Similar to the mechanisms of coral acclimatization and 
adaptation described above, there is a rapidly growing body of 
literature on the responses of corals to ocean warming (Ateweberhan et 
al., 2013; Baker et al., 2013; Bellantuono et al., 2012; Castillo et 
al., 2012; Coles and Riegl, 2013; Penin et al., 2013). These studies 
help explain the capacity for reef-building corals to acclimatize and 
adapt to ocean warming and warming-induced bleaching and suggest some 
limited capacity. However, any such capacity is highly dependent on 
species, location, habitat type, and many other factors. Available 
species-specific information on vulnerability to ocean warming and 
warming-induced bleaching, including evidence of acclimatization or 
adaptation, is provided in the Species-specific Information and 
Determination sections below.
    After considering this supplemental information in addition to that 
which was available for the proposed rule, our conclusion regarding 
ocean warming remains unchanged from the proposed rule, in that we 
consider ocean warming to be of high importance in contributing to 
extinction risk for the 65 corals in this final rule. However, we 
acknowledge that the interpretation of future ocean warming and 
warming-induced impacts to corals and coral reefs is associated with 
complexity and uncertainty, and that precise effects on individual 
species of reef-building corals are especially difficult to determine. 
The impact of ocean warming may be mediated by several factors and the 
extent to which the extinction risk of a coral species is impacted by 
ocean warming depends on its particular level of susceptibility, 
combined with its spatial and demographic characteristics in the 
context of worsening environmental conditions out to 2100, which is 
discussed in detail for each species in the Species-specific 
Information and Determinations section.

Disease (High Importance Threat, ESA Factor C)

    Disease is considered under ESA Factor C--disease or predation. In 
the proposed rule we described the threat of disease as follows. 
Disease adversely affects various coral life history events by, among 
other processes, causing adult mortality, reducing sexual and asexual 
reproductive success, and impairing colony growth. A diseased state 
results from a complex interplay of factors including the cause or 
agent (e.g., pathogen, environmental toxicant), the host, and the 
environment. All coral disease impacts are presumed to be attributable 
to infectious diseases or to poorly-described genetic defects. Coral 
disease often produces acute tissue loss. Other manifestations of 
disease in the broader sense, such as coral bleaching from ocean 
warming, are incorporated under other factors (e.g., manmade factors 
such as ocean warming as a result of climate change).
    Coral diseases are a common and significant threat affecting most 
or all coral species and regions to some degree, although the 
scientific understanding of individual disease causes in corals remains 
very poor. The incidence of coral disease appears to be expanding 
geographically in the Indo-Pacific, and there is evidence that corals 
with massive morphology species are not recovering from disease events 
in certain locations. The prevalence of disease is highly variable 
between sites and species. Increased prevalence and severity of 
diseases is correlated with increased water temperatures, which may 
correspond to increased virulence of pathogens, decreased resistance of 
hosts, or both. Moreover, the expanding coral disease threat may result 
from opportunistic pathogens that become damaging only in situations 
where the host integrity is compromised by physiological stress or 
immune suppression. Overall, there is mounting evidence that warming 
temperatures and coral bleaching responses are linked (albeit with 
mixed correlations) with increased coral disease prevalence and 
mortality. Complex aspects of temperature regimes, including winter and 
summer extremes, may influence disease outbreaks. Bleaching and coral 
abundance seem to increase the susceptibility of corals to disease 
contraction. Further, most recent research shows strong correlations 
between elevated human population density in close proximity to coral 
reefs and disease prevalence in corals.
    Although disease causes in corals remain poorly understood, some 
general patterns of biological susceptibility are beginning to emerge. 
There appear to be predictable patterns of immune capacity across coral 
families, corresponding with trade-offs with their life history traits, 
such as reproductive output and growth rate. Both Acroporidae and 
Pocilloporidae have low immunity to

[[Page 53894]]

disease. However, both of these families have intermediate to high 
reproductive outputs. Both Faviidae and Mussidae are intermediate to 
high in terms of disease immunity and reproductive output. Finally, 
while Poritidae has high immunity to disease, it has a low reproductive 
output.
    The effects of coral disease depend on exposure of the species to 
the threat, which varies spatially across the range of the species and 
temporally over time. Exposure to coral disease is moderated by 
distance of some coral habitats from the primary causes of most disease 
outbreaks, such as stressors resulting from sedimentation and nutrient 
over-enrichment. Exposure to coral disease can also be moderated by 
depth of many habitats, with deep habitats generally being less 
affected by disease outbreaks associated with stressors resulting from 
ocean warming. Disease exposure in remote areas and deep habitats 
appears to be low but gradually increasing. Exposure to coral disease 
will increase as factors that increase disease outbreaks (e.g., warming 
events) expand over time.
    As explained above, disease may be caused by threats such as ocean 
warming and bleaching, nutrients, and toxins. However, interactive 
effects between independently-arising disease and other threats are 
also important, because diseased colonies are more susceptible to the 
effects of some other threats. For example, diseased or recovering 
colonies may become more quickly stressed than healthy colonies by 
land-based sources of pollution (sedimentation, nutrients, and toxins), 
may more quickly succumb to predators, and may more easily break during 
storms or as a result of other physical impacts.
    Comments 17 and 18 discussed the importance of disease as a threat 
to corals and provided a few scientific studies (Harvell et al., 1999; 
Harvell et al., 2002; Muller and van Woesik, 2012; Rogers and Muller, 
2012) to emphasize this importance. Muller and van Woesik (2012) 
examined spatial epidemiology in the Caribbean to test if pathogens are 
contagious and spread from infected to susceptible hosts. They found no 
evidence of clustering for these diseases, so they did not follow a 
contagious disease model. They suggest the expression of coral disease 
is a two-step model: Environmental thresholds are exceeded, then those 
conditions either weaken the coral or increase the virulence of the 
pathogen (Muller and van Woesik, 2012).
    We also gathered supplemental information on the threat of disease 
since the proposed rule was published. Burge et al. (2014) summarized 
the current understanding of interactions among coral disease, elevated 
temperature, and bleaching. This supplemental information provides 
further insight of coral disease impacts at the individual level and 
the local aggregation level, and provides future predictions for the 
role of coral disease at the population level.
    At the individual level, recent studies examine both underlying 
factors and mechanistic explanations for the contraction and expansion 
of coral disease. For example, one study investigated microbial 
community dynamics in the mucus layer of corals to understand how the 
surface microbial community responds to changes in environmental 
conditions and under what circumstances it becomes vulnerable to 
overgrowth by pathogens. They found that a transient thermal anomaly 
can cause the microbial community to shift from a stable state 
dominated by antibiotic microbes to a stable state dominated by 
pathogens. Beneficial microbes may not be able to resume dominance 
after a temperature disturbance until the environment becomes 
considerably more favorable for them (Mao-Jones et al., 2010). Another 
study conducted a meta-analysis to determine whether the presence of 
particular microbial taxa correlates with the state of coral health and 
found distinct differences in the microbial taxa present in diseased 
and healthy corals (Mouchka et al., 2010). A third study investigated 
three variables commonly associated with immunity in hard and soft 
corals spanning ten families on the GBR. They found that all three 
variables (phenlyoxidase activity, size of melanin containing granular 
cells, and fluorescent protein concentrations) were significant 
predictors of susceptibility (Palmer et al., 2010). Many other studies 
have focused on bacterial or eukaryotic pathogens as the source of 
coral disease; however, a more recent study examined the role of 
viruses and determined that a specific group of viruses is associated 
with diseased Caribbean corals (Soffer et al., 2013).
    Several studies provide further evidence of disease outbreaks that 
were significantly correlated with bleaching events. The bleaching 
occurred first, then several months to a year later, there were 
significant increases in disease prevalence in bleached areas (Ban et 
al., 2013; Brandt and McManus, 2009; Bruno et al., 2007; Croquer et 
al., 2006; Croquer and Weil, 2009; Miller et al., 2009). The specific 
interactions between the two phenomena varied among disease-bleaching 
combinations. Results from one of these studies suggest the 
hypothesized relationship between bleaching and disease events may be 
weaker than previously thought, and more likely to be driven by common 
responses to environmental stressors, rather than directly facilitating 
one another.
    Ateweberhan et al. (2013) reviewed and summarized interactions 
between important threats to corals. They note that disease can 
interact not only with ocean warming and bleaching events, but may also 
be exacerbated by sedimentation, nutrients, overfishing, and 
destructive practices on coral reefs. From a broad, population-wide 
perspective, Yakob and Mumby (2011) provide an important alternative 
context in which to demonstrate that high population turnover within 
novel ecosystems (those that are different from the past and created by 
climate change) may enhance coral resistance to disease. They emphasize 
the need to move away from future projections based on historical 
trends and start to account for novel behavior of ecosystems under 
climate change.
    After considering this supplemental information in addition to that 
which was available for the proposed rule, our conclusion regarding 
disease remains unchanged from the proposed rule, in that we consider 
coral disease to be of high importance in contributing to extinction 
risk for the 65 corals in this final rule. The impact of disease may be 
mediated by several factors and the extent to which the extinction risk 
of a coral species is impacted by disease depends on its particular 
level of susceptibility, combined with its spatial and demographic 
characteristics in the context of worsening environmental conditions 
out to 2100, which is discussed in detail for each species in the 
Species-specific Information and Determinations section.

Ocean Acidification (Medium-High Importance Threat, ESA Factor E)

    Ocean acidification is considered under ESA Factor E--other natural 
or manmade factors affecting the continued existence of the species--
because the effect is a result of human activity and affects 
individuals of the coral species more so than their habitats. In the 
proposed rule we described that ocean acidification is a result of 
global climate change caused by increased GHG accumulation in the 
atmosphere. Reef-building corals produce skeletons made of the 
aragonite form of calcium carbonate; thus, reductions in aragonite 
saturation state caused by ocean acidification pose a major threat to 
these species and other

[[Page 53895]]

marine calcifiers. Ocean acidification has the potential to cause 
substantial reduction in coral calcification and reef cementation. 
Further, ocean acidification adversely affects adult growth rates and 
fecundity, fertilization, pelagic planula settlement, polyp 
development, and juvenile growth. The impacts of ocean acidification 
can lead to increased colony breakage and fragmentation and mortality. 
Based on observations in areas with naturally low pH, the effects of 
increasing ocean acidification may also include potential reductions in 
coral size, cover, diversity, and structural complexity.
    As CO2 concentrations increase in the atmosphere, more 
CO2 is absorbed by the oceans, causing lower pH and reduced 
availability of carbonate ions, which in turn results in lower 
aragonite saturation state in seawater. Because of the increase in 
CO2 and other GHGs in the atmosphere since the Industrial 
Revolution, ocean acidification has already occurred throughout the 
world's oceans, including in the Caribbean and Indo-Pacific, and is 
predicted to considerably increase between now and 2100, as described 
above in the RCP8.5 Projections section. Along with ocean warming and 
disease, we considered ocean acidification to be one of the most 
important threats posing extinction risks to coral species between now 
and the year 2100; however, individual susceptibility varies among the 
proposed species.
    Numerous laboratory and field experiments have shown a relationship 
between elevated CO2 and decreased calcification rates in 
some corals and other calcium carbonate secreting organisms. However, 
because only a few species have been tested for such effects, it is 
uncertain how most will fare in increasingly acidified oceans. In 
addition to laboratory studies, recent field studies have demonstrated 
a decline in linear growth rates of some coral species, suggesting that 
ocean acidification is already significantly reducing growth of corals 
on reefs. However, this has not been widely demonstrated across coral 
species and reef locations, suggesting species-specific effects and 
localized variability in aragonite saturation state. A potential 
secondary effect is that ocean acidification may reduce the threshold 
at which bleaching occurs. Overall, the best available information 
demonstrates that most corals exhibit declining calcification rates 
with rising CO2 concentrations, declining pH, and declining 
aragonite saturation state, although the rate and mode of decline can 
vary among species. Recent studies also discuss the physiological 
effects of ocean acidification on corals and their responses. Corals 
are able to regulate pH within their tissues, maintaining higher pH 
values in their tissues than the pH of surrounding waters. This is an 
important mechanism in naturally highly-fluctuating environments (e.g., 
many backreef pools have diurnally fluctuating pH) and suggests that 
corals have some adaptive capacity to acidification. However, as with 
ocean warming, there is high uncertainty as to whether corals will be 
able to adapt quickly enough to the projected changes in aragonite 
saturation state.
    In addition to the direct effects on coral calcification and 
growth, ocean acidification may also affect coral recruitment, reef 
cementation, and other important reef-building species like crustose 
coralline algae. Studies suggest that the low pH associated with ocean 
acidification may impact coral larvae in several ways, including 
reduced survival and recruitment. Ocean acidification may influence 
settlement of coral larvae on coral reefs more by indirect alterations 
of the benthic community, which provides settlement cues, than by 
direct physiological disruption. A major potential impact from ocean 
acidification is a reduction in the structural stability of corals and 
reefs, which results both from increases in bioerosion and decreases in 
reef cementation. As atmospheric CO2 rises globally, reef-
building corals are expected to calcify more slowly and become more 
fragile. Declining growth rates of crustose coralline algae may 
facilitate increased bioerosion of coral reefs from ocean 
acidification. Studies demonstrate that ocean acidification will likely 
have a great impact on corals and reef communities by affecting 
community composition and dynamics, exacerbating the effects of disease 
and other stressors (e.g., temperature), contributing to habitat loss, 
and affecting symbiont function. Some studies have found that an 
atmospheric CO2-level twice as high as pre-industrial levels 
will start to dissolve coral reefs; this level could be reached as 
early as the middle of this century. Further, the rate of acidification 
may be an order of magnitude faster than what occurred 55 million years 
ago during the Paleocene-Eocene Thermal Maximum (i.e., the period in 
which global temperatures rose 5 to 9 [deg]C, providing a context in 
which to understand climate change).
    While CO2 levels in the surface waters of the ocean are 
generally in equilibrium with the lower atmosphere, there can be 
considerable variability in seawater pH across reef-building coral 
habitats, resulting in colonies of a species experiencing high spatial 
variability in exposure to ocean acidification. The spatial variability 
in seawater pH occurs from reef to global scales, driven by numerous 
physical and biological characteristics and processes, including: 
Seawater temperature; proximity to land-based runoff and seeps; 
proximity to sources of oceanic CO2; salinity; nutrients; 
photosynthesis; and respiration. In cooler waters, CO2 
absorption is higher, driving pH and aragonite saturation state lower, 
thus relatively cool coral habitats are more susceptible to 
acidification, such as those at higher latitudes, in upwelling areas, 
and in deeper environments. On coral reefs, wave and wind-induced 
mixing typically maintain roughly similar temperatures in the shallow 
photic zone preferred by most reef-building corals, thus the deeper 
environments that are more susceptible to acidification are generally 
below this photic zone.
    Land-based runoff decreases salinity and increases nutrients, both 
of which can raise pH. Local sources of oceanic CO2 like 
upwelling and volcanic seeps lower pH. Photosynthesis in algae and 
seagrass beds draws down CO2, raising pH. High variability 
over various time-scales is produced by numerous processes, including 
diurnal cycles of photosynthesis and respiration, seasonal variability 
in seawater temperatures, and decadal cycles in upwelling. Temporal 
variability in pH can be very high diurnally in highly-fluctuating or 
semi-enclosed habitats such as reef flats and back-reef pools, due to 
high photosynthesis during the day (pH goes up) and high respiration 
during the night (pH goes down). In fact, pH fluctuations during one 
24-hr period in such reef-building coral habitats can exceed the 
magnitude of change expected by 2100 in open ocean subtropical and 
tropical waters. As with spatial variability in exposure to ocean 
warming, temporal variability in exposure to ocean acidification is a 
combination of high variability over short time-scales together with 
long-term increases. While exposure of the proposed coral species to 
ocean acidification varies greatly both spatially and temporally, it is 
expected to increase for all species across their ranges between now 
and 2100.
    Ocean acidification likely interacts with other threats, especially 
considering that ocean acidification is expected to continue to 
increase over the foreseeable future. For example, ocean acidification 
may reduce the threshold at which bleaching occurs, increasing the 
threat posed by ocean warming. One of the key impacts of ocean 
acidification is reduced

[[Page 53896]]

calcification, resulting in reduced skeletal growth and skeletal 
density, which may lead to numerous interactive effects with other 
threats. Reduced skeletal growth compromises the ability of coral 
colonies to compete for space against algae, which grows more quickly 
as nutrient over-enrichment increases, especially if not held in check 
by herbivores. Reduced skeletal density weakens coral skeletons, 
resulting in greater colony breakage from natural and human-induced 
physical damage.
    As discussed in Comments 18-21, we received numerous comments 
related to the threat to corals from ocean acidification including: (1) 
The overview and future projections of ocean acidification; (2) 
variability in ocean acidification; and (3) specific effects of ocean 
acidification on reef-building corals.
    Comment 17 stated that we oversimplified the complexity and 
variability in the future projections of ocean acidification, and 
criticized our reliance on AR4 as the basis for our threat evaluation. 
In the proposed rule, we acknowledged the uncertainty associated with 
projections of ocean acidification from global climate change. However, 
while there are many sources of uncertainty in climate change 
projections, and likewise for ocean acidification, the ocean 
acidification projections in AR4 and AR5's WGI represent the best 
available information. The proposed rule and supporting documents 
assumed that AR4's highest-emission scenario A1F1 was the most likely 
to occur. Now that AR5's WGI is available, we consider the most 
impactful pathway to coral is WGI's RCP8.5, which includes ocean 
acidification projections. These projections are described above in the 
RCP8.5 Projections section, along with two independent analyses of the 
effects of ocean acidification projections in RCP8.5 on coral reefs in 
the 21st century. As noted in the RCP8.5 Projections section, there is 
uncertainty in these ocean acidification projections for coral reefs.
    Comment 18 specifically cites Manzello et al. (2012) and Palacios 
and Zimmerman (2012; 2007) to illustrate that variability in ocean 
acidification on coral reefs can be buffered by local and regional 
biogeochemical processes within seagrass beds. Additionally, 
biogeochemical processes within coral reef communities (Andersson et 
al., 2013) may buffer the effects of decreasing pH. Other scientific 
studies identify mechanisms that can exacerbate changes in seawater pH 
around coral reefs from ocean acidification, such as diurnal 
variability that can amplify CO2 in seawater around coral 
reefs (Shaw et al., 2013). On larger scales, a recent study 
demonstrated that some coastal areas of the Gulf of Mexico and South 
Atlantic were buffered against ocean acidification because of the input 
of fresh, alkaline surface waters carrying dissolved inorganic carbon 
(Wang et al., 2013). Variability in ocean acidification at basin and 
global scales is influenced largely by upwelling and latitude, with 
more acidification in areas of high upwelling and lower temperatures. 
The interaction of ocean acidification with ocean warming produces 
basin-level patterns of higher and lower habitat suitability for reef-
building corals (Couce et al., 2013b; van Hooidonk et al., 2013a; Yara 
et al., 2012).
    Comments 19 and 20 underscore specific effects to corals from ocean 
acidification identified in the proposed rule, including: (1) Effects 
on reef accretion; (2) effects on larvae and juvenile corals; (3) 
interactive effects with other environmental variables; and (4) 
miscellaneous effects. Recent research identifies impacts of ocean 
acidification on reef accretion due to reduced coral calcification 
(Chan and Connolly, 2013) and impacts on crustose coralline algae 
(Doropoulos and Diaz-Pulido, 2013). Recent research has also found that 
impacts of ocean acidification on brooded larvae of Pocillopora 
damicornis were higher when the larvae were released earlier (Cumbo et 
al., 2013) and that nutritionally replete juvenile corals were less 
susceptible to ocean acidification than nutritionally deprived 
juveniles (Drenkard et al., 2013).
    Many recent studies have investigated the interactive effects of 
ocean acidification with other environmental variables. The opposing 
effects of ocean warming and ocean acidification were discussed in a 
study that demonstrated low light conditions can exacerbate ocean 
acidification effects. Low-light conditions can provide a refuge for 
reef-building corals from thermal and light stress, but this study 
suggests that lower light availability will potentially increase the 
susceptibility of key coral species to ocean acidification (Suggett et 
al., 2013). Another study predicts that increasing storms predicted by 
climate change, together with ocean acidification, are likely to 
increase collapse of table corals (Madin et al., 2012). Salinity 
extremes on a nearshore coral community did not affect the sensitivity 
of reef-building corals to ocean acidification (Okazaki et al., 2013). 
Finally, several studies have investigated the simultaneous effects of 
ocean warming and ocean acidification, most of which have found harmful 
synergistic effects (Ateweberhan et al., 2013; Dove et al., 2013; 
Kroeker et al., 2013), but not all (Wall et al., 2013). However, 
impacts of ocean acidification are more rapid in cool water, such as in 
mesophotic habitat (Cerrano et al., 2013) and temperate areas (Yara et 
al., 2012).
    Several other recent papers also provide information on the impacts 
of ocean acidification on reef-building corals. A study of the effects 
of ocean acidification on primary polyps with and without zooxanthellae 
found that polyps with zooxanthellae had higher tolerance to ocean 
acidification, suggesting that coral species that acquire symbionts 
from the environment will be more vulnerable to ocean acidification 
than corals that maternally acquire symbionts (i.e., brooding species; 
Ohki et al., 2013). A study of Porites corals at a field site with 
naturally low pH found that the corals were not able to acclimatize 
enough to prevent the impacts of local ocean acidification on their 
skeletal growth and development, despite spending their entire lifespan 
in low pH seawater (Crook et al., 2013). A study of the effects of 
ocean acidification on different coral species in different 
environments found that effects were highly species-dependent, and 
furthermore, that effects within a species depended on the environment 
(Kroeker et al., 2013).
    After considering this supplemental information in addition to that 
which was available for the proposed rule, our conclusion regarding 
ocean acidification remains unchanged from the proposed rule, in that 
we consider ocean acidification to be of medium-high importance in 
contributing to extinction risk for the 65 corals in this final rule. 
However, we acknowledge that the interpretation of future ocean 
acidification and acidification-induced impacts to corals and coral 
reefs is associated with complexity and uncertainty and that the 
effects on individual species of reef-building corals are especially 
difficult to determine. The impact of ocean acidification may be 
mediated by several factors and the extent to which the extinction risk 
of a coral species is impacted by ocean acidification depends on its 
particular level of susceptibility, combined with its spatial and 
demographic characteristics in the context of worsening environmental 
conditions out to 2100, which is discussed in detail for each species 
in the Species-specific Information and Determinations section.

[[Page 53897]]

Trophic Effects of Fishing (Medium Importance Threat, ESA Factor A)

    Trophic effects of fishing are considered under ESA Factor A--the 
present or threatened destruction, modification, or curtailment of its 
habitat or range--because the main effect of concern is to limit 
availability of habitat for corals. In the proposed rule we described 
the threat of the trophic effects of reef fishing as follows. Fishing, 
particularly overfishing, can have large scale, long-term ecosystem-
level effects that can change ecosystem structure from coral-dominated 
reefs to algal-dominated reefs (``phase shifts''). Even fishing 
pressure that doesn't rise to the level of overfishing potentially can 
alter trophic interactions that are important in structuring coral reef 
ecosystems. These trophic interactions include reducing population 
abundance of herbivorous fish species that control algal growth, 
limiting the size structure of fish populations, reducing species 
richness of herbivorous fish, and releasing corallivores from predator 
control. Thus, an important aspect of maintaining resilience in coral 
reef ecosystems is to sustain populations of herbivores, especially the 
larger scarine herbivorous wrasses such as parrotfish.
    On topographically complex reefs, population densities can average 
well over a million herbivorous fishes per km\2\, and standing stocks 
can reach 45 metric tons per km\2\. In the Caribbean, parrotfishes can 
graze at rates of more than 150,000 bites per square meter per day, and 
thereby remove up to 90-100 percent of the daily primary production 
(e.g., algae). Under these conditions of topographic complexity with 
substantial populations of herbivorous fishes, as long as the cover of 
living coral is high and resistant to mortality from environmental 
changes, it is very unlikely that the algae will take over and dominate 
the substrate. However, if herbivorous fish populations, particularly 
large-bodied parrotfish, are heavily fished and a major mortality of 
coral colonies occurs, then algae can grow rapidly and prevent the 
recovery of the coral population. The ecosystem can then collapse into 
an alternative stable state, a persistent phase shift in which algae 
replace corals as the dominant reef species. Although algae can have 
negative effects on adult coral colonies (e.g., overgrowth, bleaching 
from toxic compounds), the ecosystem-level effects of algae are 
primarily from inhibited coral recruitment. Filamentous algae can 
prevent the colonization of the substrate by planula larvae by creating 
sediment traps that obstruct access to a hard substrate for attachment. 
Additionally, macroalgae can suppress the successful colonization of 
the substrate by corals through occupation of the available space, 
shading, abrasion, chemical poisoning, and infection with bacterial 
disease.
    Overfishing can have further impacts on coral mortality via trophic 
cascades. In general larger fish are targeted, resulting in fish 
populations of small individuals. For parrotfishes, the effect of 
grazing by individuals greater than 20 cm in length is substantially 
greater than that by smaller fish. Up to 75 individual parrotfishes 
with lengths of about 15 cm are necessary to have the same reduction in 
algae and promotion of coral recruitment as a single individual 35 cm 
in length. Species richness of the herbivorous fish population is also 
very beneficial to maintaining available substrate potentially leading 
to enhanced coral populations. Because of differences in their feeding 
behaviors, several species of herbivorous fishes with complementary 
feeding behaviors can have a substantially greater positive effect than 
a similar biomass of a single species on reducing the standing stock of 
macroalgae, of increasing the cover of crustose coralline algae, and 
increasing live coral cover.
    Exposure to the trophic effects of fishing in the Caribbean may be 
moderated by distance of some coral habitats from fishing effort. 
Exposure to the trophic effects of fishing in the Indo-Pacific is 
likely more moderated by distance than in the Caribbean, due to a 
greater proportion of reef-building coral habitats located in remote 
areas away from fishing effort. Exposure to the trophic effects of reef 
fishing may also moderated by depth of many habitats in both regions, 
but again more so in the Indo-Pacific than in the Caribbean. Deep 
habitats are generally less affected by the trophic effects of fishing, 
especially in the Indo-Pacific. Exposure to the trophic effects of 
fishing will increase as the human population increases over time.
    The trophic effects of fishing are likely to interact with many 
other threats, especially considering that fishing impacts are likely 
to increase within the ranges of many of the proposed corals over the 
foreseeable future. For example, when carnivorous fishes are 
overfished, corallivore populations may increase, resulting in greater 
predation on corals. Further, overfishing appears to increase the 
frequency of coral disease. Fishing activity usually targets the larger 
apex predators. When predators are removed, corallivorous butterfly 
fishes become more abundant and can transmit disease from one coral 
colony to another as they transit and consume from each coral colony. 
With increasing abundance, they transmit disease to higher proportions 
of the corals within the population.
    Comments 21-23 focused on the following aspects of the trophic 
effects of reef fishing: (1) The importance of the threat to coral 
reefs; (2) higher importance localized threats; and (3) consideration 
of human demography.
    Comment 21 highlighted Keller et al. (2009), which provides 
additional support for the importance herbivores play in the 
maintenance of recruitment habitat. Further, recent information shows 
that one of the most detrimental effects of unsustainable fishing 
pressure is the alteration of trophic interactions that are 
particularly important in structuring coral reef ecosystems (Jackson et 
al., 2012; Jackson et al., 2014; Ruppert et al., 2013). These trophic 
interactions include reducing population abundance of herbivorous fish 
species that control algal growth, limiting the size structure of fish 
populations, reducing species richness of herbivorous fish, and 
releasing corallivores from predator control. Thus, an important aspect 
of maintaining resilience in coral reef ecosystems is to sustain 
functional populations of herbivores, especially the larger parrotfish 
and other key functional herbivorous fish (Hughes et al., 2010; Jackson 
et al., 2012; Jackson et al., 2014; Kennedy et al., 2013). Further, 
Jackson et al. (2014) considers overfishing (associated with high human 
densities) to be one of the major causes of the region-wide decline in 
Caribbean corals while acknowledging that climate threats are likely to 
be major sources of mortality in the future. In addition to direct 
overfishing of primary consumers such as parrotfish, recent studies 
found that overfishing of top reef predators such as sharks and other 
predatory fish, such as large groupers in the Caribbean, can have an 
impact that cascades down the food chain, potentially contributing to 
mesopredator release, and ultimately altering the numbers of primary 
consumers available to control algal growth (Jackson et al., 2012; 
Jackson et al., 2014; Ruppert et al., 2013).
    After considering this supplemental information in addition to that 
which was available for the proposed rule, our conclusion regarding the 
trophic effects of fishing remains unchanged from the proposed rule. 
Trophic effects of fishing are a medium importance threat in assessing 
global extinction risk for the 65 corals in this final rule. Because 
the main effect of trophic effects of reef

[[Page 53898]]

fishing is habitat alteration, there are no species-specific levels of 
exposure and susceptibility. However, the extent to which an individual 
species' recruitment is affected is discussed in more detail in the 
Species-specific Information and Determinations section, when species-
specific information is available.

Sedimentation (Low-Medium Importance Threat, ESA Factors A and E)

    Sedimentation is considered under ESA Factor A--the present or 
threatened destruction, modification, or curtailment of its habitat or 
range--and ESA Factor E--other natural or manmade factors affecting the 
continued existence of the species--because the effect of the threat, 
resulting from human activity, is both to limit the availability of 
habitat for corals and to directly impact individuals of coral species. 
In the proposed rule we described the threat of sedimentation as 
follows. Human activities in coastal and inland watersheds introduce 
sediment into the ocean by a variety of mechanisms, including river 
discharge, surface runoff, groundwater seeps, and atmospheric 
deposition. Humans also introduce sewage into coastal waters through 
direct discharge, treatment plants, and septic leakage. Elevated 
sediment levels are generated by poor land use practices and coastal 
and nearshore construction.
    The most common direct effect of sedimentation is deposition of 
sediment on coral surfaces as sediment settles out from the water 
column. Corals with certain morphologies (e.g., mounding) can passively 
reject settling sediments. In addition, corals can actively displace 
sediment by ciliary action or mucous production, both of which require 
energetic expenditures. Corals with large calices (skeletal component 
that holds the polyp) tend to be better at actively rejecting sediment. 
Some coral species can tolerate complete burial for several days. 
Corals that are unsuccessful in removing sediment will be smothered and 
die. Sediment can also induce sublethal effects, such as reductions in 
tissue thickness, polyp swelling, zooxanthellae loss, and excess mucus 
production. In addition, suspended sediment can reduce the amount of 
light in the water column, making less energy available for coral 
photosynthesis and growth. Sedimentation also impedes fertilization of 
spawned gametes and reduces larval settlement and survival of recruits 
and juveniles.
    Although it is difficult to quantitatively predict the extinction 
risk that sedimentation poses to the corals under consideration, human 
activity has resulted in quantifiable increases in sediment inputs in 
some reef areas. Continued increases in coastal human populations 
combined with poor land use and nearshore development practices will 
likely increase sediment delivery to reef systems. Nearshore sediment 
levels will also likely increase with sea-level rise. Greater 
inundation of reef flats can erode soil at the shoreline and resuspend 
lagoon deposits, producing greater sediment transport and potentially 
leading to leeward reefs being flooded with turbid lagoon waters or 
buried by off-bank sediment transport. Sediment stress and turbidity 
also can induce bleaching, although some corals may be more tolerant of 
elevated short-term levels of sedimentation.
    Exposure to sedimentation can be moderated by distance of some 
coral habitats from areas where sedimentation is chronically or 
sporadically heavy, resulting in some habitats being unaffected or very 
lightly affected by sedimentation. Exposure to sedimentation for 
particular species may also be moderated by depth of habitats. Exposure 
to sedimentation is expected to increase as human activities that 
produce sedimentation expand over time.
    Sedimentation is also likely to interact with many other threats, 
such as other land-based sources of pollution and warming-induced 
bleaching, especially considering that sedimentation is likely to 
increase across the ranges of many of the 65 species over the 
foreseeable future. For example, when coral communities that are 
chronically affected by sedimentation experience a warming-induced 
bleaching event, a disease outbreak, or a toxic spill, the consequences 
for those corals can be much more severe than in communities not 
affected by sedimentation.
    Comment 25 underscored the importance of sedimentation as a 
considerable local threat to corals, and highlighted the potential of 
sedimentation to interact and potentially exacerbate other threats. A 
few commenters provided references (Bonkosky et al., 2009; 
Hern[aacute]ndez- Delgado et al., 2012; Hernandez-Delgado et al., 2011) 
that discussed sedimentation as a threat to corals.
    We also gathered supplemental studies on the threat of 
sedimentation since the proposed rule was published. Three points in 
particular from the proposed rule were affirmed by the supplemental 
studies. Sedimentation can have interactive effects with other 
stressors including disease and climate factors such as bleaching 
susceptibility and reduced calcification (Ateweberhan et al., 2013; 
Suggett et al., 2013). MPAs provide little protection against indirect 
stressors like sedimentation from upland activities (Halpern et al., 
2013). The effects of sedimentation can be variable for different coral 
species and may depend on other environmental conditions (Blakeway et 
al., 2013; Suggett et al., 2013).
    After considering this supplemental information in addition to that 
which was available for the proposed rule, our conclusion regarding 
sedimentation remains unchanged from the proposed rule. Sedimentation 
is a low to medium importance threat in assessing global extinction 
risk for the 65 corals in this final rule. The impact of sedimentation 
may be mediated by several factors and the extent to which the 
extinction risk of a coral species is impacted by sedimentation depends 
on its particular level of susceptibility combined with it spatial and 
demographic characteristics in the context of worsening environmental 
conditions out to 2100, which is considered for each species in the 
Species-specific Information and Determinations section.

Nutrients (Low-Medium Importance Threat, ESA Factors A and E)

    Nutrient enrichment is considered under ESA Factor A--the present 
or threatened destruction, modification, or curtailment of its habitat 
or range--and ESA Factor E--other natural or manmade factors affecting 
the continued existence of the species--because the effect of the 
threat, resulting from human activity, is both to limit the 
availability of habitat for corals and directly impact individuals of 
coral species. In the proposed rule we described the threat of nutrient 
over-enrichment as follows. Elevated nutrients affect corals through 
two main mechanisms: Direct impacts on coral physiology and indirect 
effects through nutrient-stimulation of other community components 
(e.g., macroalgal turfs and seaweeds, and filter feeders) that compete 
with corals for space on the reef. Increased nutrients can decrease 
calcification; however, nutrients may also enhance linear extension, 
while reducing skeletal density. Either condition results in corals 
that are more prone to breakage or erosion, but individual species do 
have varying tolerances to increased nutrients. The main vectors of 
anthropogenic nutrients are point-source discharges (such as rivers or 
sewage outfalls) and surface runoff from modified watersheds. Natural 
processes,

[[Page 53899]]

such as in situ nitrogen fixation and delivery of nutrient-rich deep 
water by internal waves and upwelling also bring nutrients to coral 
reefs.
    Exposure to nutrients can be moderated by distance of some coral 
habitats from areas where nutrients are chronically or sporadically 
heavy (e.g., heavily populated areas). However, nutrient over-
enrichment can still result from sparsely populated areas; and these 
nutrients can be quickly transported large distances. Therefore, 
distance is less of a moderating factor for nutrients than for 
sedimentation. Similarly, although nutrient exposure may also be 
moderated by the depth of some habitats, nutrient impacts extend deeper 
than sedimentation impacts. Exposure to nutrients is expected to 
increase as human activities that produce nutrients expand over time.
    Nutrients are likely to interact with many other threats, 
especially considering that nutrient over-enrichment is likely to 
increase across the ranges of many of the 65 corals over the 
foreseeable future. For example, when coral communities that are 
chronically affected by nutrients experience a warming-induced 
bleaching event, a disease outbreak, or a toxic spill, the consequences 
for corals can be much more severe than in communities not affected by 
nutrients.
    Comment 26 supported and reiterated the effects nutrients can have 
on corals. Some of the individual commenters provided studies (Bonkosky 
et al., 2009; Connolly et al., 2012; Cunning and Baker, 2013; Fabricius 
et al., 2013; Hernandez-Delgado et al., 2011; Hern[aacute]ndez-Delgado 
et al., 2008; M[eacute]ndez-L[aacute]zaro et al., 2012; Wiedenmann et 
al., 2013) to reinforce their support. Bonkosky et al. (2009) provided 
further evidence that elevated turbidity and nutrient enrichment from 
human waste discharge has an extensive impact on coral reef ecosystems. 
In response to contradictory results from other studies as to whether 
nutrients increase thermal stress or increase resistance to higher 
temperature for corals, Fabricius et al. (2013) exposed corals to both 
elevated nutrients and heat stress. They found higher mortality 
occurred in the elevated nutrient-heat stress treatments versus heat-
stressed alone and controls. Wiedenmann et al. (2013) found that 
unfavorable ratios of dissolved inorganic nutrients in the water column 
led to phosphate starvation of symbiotic algae, reducing thermal 
tolerance. Cunning and Baker (2013) found higher nutrient loads can 
lead to higher densities of symbionts, and corals with higher densities 
of symbionts were more susceptible to bleaching.
    We also gathered supplemental information on how elevated nutrients 
interact with other threats, including coral bleaching and disease. One 
study tested the interactive effects of nutrient loading with both 
bleaching and disease and found that coral disease prevalence and 
severity as well as coral bleaching were increased in nutrient enriched 
plots (Vega Thurber et al., 2013). Ateweberhan et al. (2013) note that 
most studies on the subject of nutrient enrichment and high 
temperatures also present evidence of negative effects on calcification 
due to higher nutrient levels, although both positive and negative 
effects have been reported. Nutrient enrichment can also interact with 
the threat of coral disease by encouraging the proliferation of 
disease-causing microorganisms and bioeroders, such as boring sponges, 
and intensifying the growth of fleshy macroalgae that harbor and spread 
coral diseases (Ateweberhan et al., 2013; Vega Thurber et al., 2013).
    After considering this supplemental information in addition to that 
which was available for the proposed rule, our conclusion regarding 
nutrient over-enrichment remains unchanged from the proposed rule. 
Nutrients are a low to medium importance threat in assessing global 
extinction risk for the 65 corals in this final rule. The impact of 
elevated nutrients may be mediated by several factors and the extent to 
which the extinction risk of a coral species is impacted by nutrient 
enrichment depends on its particular level of susceptibility, combined 
with its spatial and demographic characteristics in the context of 
worsening environmental conditions out to 2100, which is considered for 
each species in the Species-specific Information and Determinations 
section.

Sea-Level Rise (Low-Medium Threat, ESA Factor A)

    Sea-level rise is considered under ESA Factor A--the present or 
threatened destruction, modification, or curtailment of its habitat or 
range--because the effect of the threat is to the availability of 
corals' habitat and not directly to the species themselves. In the 
proposed rule we described the threat of sea-level rise as follows. The 
effects of sea-level rise may act on various coral life history events, 
including larval settlement, polyp development, and juvenile growth, 
and can contribute to adult mortality and colony fragmentation, mostly 
due to increased sedimentation and decreased water quality (reduced 
light availability) caused by coastal inundation. The best available 
information suggests that sea level will continue to rise due to 
thermal expansion and the melting of land and sea ice. Theoretically, 
any rise in sea-level could potentially provide additional habitat for 
corals living near the sea surface. Many corals that inhabit the 
relatively narrow zone near the ocean surface have rapid growth rates 
when healthy, which allowed them to keep up with sea-level rise during 
the past periods of rapid climate change associated with deglaciation 
and warming. However, depending on the rate and amount of sea-level 
rise, rapid rises can lead to reef drowning. Rapid rises in sea level 
could affect many of the proposed coral species by both submerging them 
below their common depth range and, more likely, by degrading water 
quality through coastal erosion and potentially severe sedimentation or 
enlargement of lagoons and shelf areas. Rising sea level is likely to 
cause mixed responses in the 65 corals depending on their depth 
preferences, sedimentation tolerances, growth rates, and the nearshore 
topography. Reductions in growth rate due to local stressors, 
bleaching, infectious disease, and ocean acidification may prevent the 
species from keeping up with sea-level rise (i.e., from growing at a 
rate that will allow them to continue to occupy their preferred depth 
range despite sea-level rise).
    The rate and amount of future sea-level rise remains uncertain. 
Until the past few years, sea-level rise was predicted to be in the 
range of only about one half meter by 2100. However, more recent 
estimated rates are higher, based upon evidence that the Greenland and 
Antarctic ice sheets are much more vulnerable than previously thought. 
While there is large variability in predictions of sea-level rise, AR4 
likely underestimated the rates under all scenarios.
    Fast-growing branching corals were able to keep up with the first 3 
m of sea-level rise during the warming that led to the last 
interglacial period. However, whether the 65 corals in this final rule 
will be able to survive 3 m or more of future sea-level rise will 
depend on whether growth rates are reduced as a result of other risk 
factors, such as local environmental stressors, bleaching, infectious 
disease, and ocean acidification. Additionally, lack of suitable new 
habitat, limited success in sexual recruitment, coastal runoff, and 
coastal hardening will compound some corals' ability to survive rapid 
sea-level rise.

[[Page 53900]]

    This threat is expected to disproportionately affect shallow areas 
adjacent to degraded coastlines, as ocean inundation results in higher 
levels of sedimentation from the newly-inundated coastlines to the 
shallow areas. Exposure to sea-level rise will be moderated by 
horizontal and vertical distances of reef-building coral habitats from 
inundated, degraded coastlines. Exposure to sea-level rise will 
increase over time as the rate of rise increases.
    Sea-level rise is likely to interact with other threats, especially 
considering that sea-level rise is likely to increase across the ranges 
of the 65 corals over the foreseeable future. In particular, the 
inundation of developed areas (e.g., urban and agricultural areas) and 
other areas where shoreline sediments are easily eroded by sea-level 
rise is likely to degrade water quality of adjacent coral habitat 
through increased sediment and nutrient runoff and the potential 
release of toxic contamination.
    Comment 27 supported the Consensus Statement on Climate Change and 
Coral Reefs, which specifies that sea-levels have already risen and 
that future rising sea-levels will be accompanied by increased 
sedimentation levels. We received no additional supplemental 
information on this threat.
    We also gathered supplemental information to update the analysis 
presented in the proposed rule. In the proposed rule, we noted that AR4 
likely underestimated rates of projected sea-level rise. AR5's WGI 
represents a substantial advance from AR4. The first section of WGI 
considers observations of climate system change, which refers to 
descriptions of past climate patterns. WGI concludes it is virtually 
certain that the global mean sea level rose by 19 cm from 1901 to 2010. 
The anthropogenic ocean warming observed since the 1970s has 
contributed to global sea-level rise over this period through ice 
melting and thermal expansion. Projections for future sea-level-rise in 
RCP8.5 for the period 2081 to 2100 are 0.53 to 0.97 meter higher than 
the period 1986 to 2005. In addition, WGI concluded that it is 
virtually certain that global mean sea-level rise will continue beyond 
2100. WGI also reported that it is very likely that in the twenty-first 
century and beyond, sea-level change will have a strong regional 
pattern (IPCC, 2013).
    After considering this supplemental information in addition to that 
which was previously available, our conclusion regarding sea-level rise 
remains unchanged from the proposed rule. Sea-level rise is a low to 
medium importance threat in assessing global extinction risk for the 65 
corals in this final rule. The impact of sea-level-rise may be mediated 
by some factors and the extent to which the extinction risk of a coral 
species is impacted by sea-level-rise depends on its particular level 
of susceptibility, combined with its spatial and demographic 
characteristics in the context of worsening environmental conditions 
out to 2100, which is considered for each species in the Species-
specific Information and Determinations section.

Predation (Low Threat, ESA Factor C)

    Predation is considered under ESA Factor C--disease or predation. 
In the proposed rule we described the threat of predation as follows. 
Predation on some coral genera by many corallivorous species of fish 
and invertebrates (e.g., snails and seastars) is a chronic, though 
occasionally acute, energy drain. It is a threat that has been 
identified for most coral life stages. Thus, predation factored into 
the extinction risk analysis for each of the 65 corals. Numerous 
studies have documented the quantitative impact of predation by various 
taxa on coral tissue and skeleton. Predators can indirectly affect the 
distribution of corals by preferentially consuming faster-growing coral 
species, thus allowing slower-growing corals to compete for space on 
the reef. The most notable example of predation impacts in the Indo-
Pacific are from large aggregations or outbreaks of crown-of-thorns 
seastar. The specific cause of crown-of-thorns seastar outbreaks is 
unknown. Crown-of-thorns seastar can reduce living coral cover to less 
than one percent during outbreaks, changing coral community structure, 
promoting algal colonization, and affecting fish population dynamics.
    Exposure to predation by corallivores is moderated by presence of 
predators of the corallivores. For example, corallivorous reef fish 
prey on corals, and piscivorous reef fish and sharks prey on the 
corallivores; thus, high abundances of piscivorous reef fish and sharks 
moderate coral predation. Abundances of piscivorous reef fish and 
sharks vary spatially because of different ecological conditions and 
human exploitation levels. Exposure to predation is also moderated by 
distance from physical conditions that allow corallivore populations to 
grow. For example, in the Indo-Pacific, high nutrient runoff from 
continents and high islands improves reproductive conditions for crown-
of-thorns seastar, thus coral predation by crown-of-thorns seastar is 
moderated by distance from such conditions. Predation can also be 
moderated by depth of many habitats because abundances of many 
corallivorous species decline with depth. Exposure to predation can 
increase over time as conditions change, but may be moderated by 
distance and depth for certain species, which depends upon the 
distribution and abundances of the species.
    Predation of coral colonies can increase the likelihood of the 
colonies being infected by disease, and likewise diseased colonies may 
be more likely to be preyed upon. There are likely other examples of 
cumulative and interactive effects of predation with other threats to 
corals.
    Comment 28 suggested predation and exposure values for some 
individual species, but did not provide supplemental information on the 
threat. We also gathered supplemental information that supports and 
reiterates the analysis presented in the proposed rule. Bonaldo et al. 
(2011) documented spatial and temporal variation in coral predation by 
parrotfishes on the Great Barrier Reef. Lenihan et al. (2011) assessed 
the degree to which the performance of recently recruited branching 
corals was influenced by several factors, including corallivory. They 
found that partial predation by corallivorous fishes is an important 
but habitat-modulated constraint for branching corals and, overall, 
corallivory had variable effects on corals of different genera. Last, 
De'ath et al. (2012) documented that 42 percent of the decline in coral 
cover on the GBR is attributable to crown-of-thorns seastar predation.
    After considering this supplemental information in addition to that 
which was available for the proposed rule, our conclusion regarding 
predation remains unchanged from the proposed rule. Predation is a low 
importance threat in assessing global extinction risk for the 65 corals 
in this final rule. The impact of predation may be mediated by several 
factors and the extent to which the extinction risk of a coral species 
is impacted by predation depends on its particular level of 
susceptibility combined with its spatial and demographic 
characteristics in the context of worsening environmental conditions 
out to 2100, which is considered for each species in the Species-
specific Information and Determinations section.

Collection and Trade (Low Threat, ESA Factor B)

    Collection and trade is considered under ESA Factor B--
overutilization for commercial, recreational, scientific, or 
educational purposes. In the proposed rule, we described the threat of 
collection and trade as follows.

[[Page 53901]]

Globally, 1.5 million live stony coral colonies are reported to be 
collected from at least 45 countries each year, with the United States 
consuming the largest portion of live corals (64 percent) and live rock 
(95 percent) for the aquarium trade. The imports of live corals taken 
directly from coral reefs (not from aquaculture) increased by 600 
percent between 1988 and 2007, while the global trade in live coral 
increased by nearly 1,500 percent. Harvest of stony corals is usually 
highly destructive, and results in removing and discarding large 
amounts of live coral that go unsold and damaging reef habitats around 
live corals. While collection is a highly spatially-focused impact, it 
can result in significant impacts and was considered to contribute to 
individual species' extinction risk. However, we ultimately ranked this 
threat as low overall because of species-specific factors (i.e., some 
species are preferentially affected) as well as distance and depth 
factors that create barriers to human access.
    As described in Comments 29 and 30, we received a significant 
amount of supplemental information via public comments and gathered 
supplemental information on three aspects of the threat of collection 
and trade on reef-building corals and coral reef ecosystems: (1) Wild 
collection of corals, including information about the physical and 
ecological impacts of wild collection of coral colonies and/or 
fragments from their natural habitats; (2) captive culture including 
information regarding the development of mariculture and aquaculture 
operations, as well as the role of home aquaria as they relate to 
trade, including all commercial, recreational, and educational coral-
raising operations in marine environments as well as in captivity; and 
(3) the global marine ornamental trade industry, including detailed 
information regarding trade of both live and dead corals and other 
coral reef wildlife.
    For the purposes of this final rule, collection and trade refers to 
the physical process of taking corals from their natural habitat on 
coral reefs for the purpose of sale in the ornamental trade industry. 
We define wild collection as the physical removal or capture of coral 
colonies, fragments, and polyps from their natural habitat. This 
section also discusses the use of captive breeding techniques via 
aquaculture and mariculture for the purposes of trade. Captive culture 
techniques are increasingly used to supply the aquarium trade industry 
and potentially reduce the amount of corals collected from the wild to 
meet demand (Thornhill, 2012; Wood et al., 2012). We define aquaculture 
as the land-based (`ex situ') propagation or grow out of corals. 
Examples of this include corals grown in home aquaria or terrestrial 
coral farms. We define mariculture as the ocean-based (`in situ') 
propagation or grow out of corals. Examples of this include corals 
grown in coral farms and nursery areas in marine environments. The 
phrase ``captive culture'' is used interchangeably to refer to captive 
breeding of corals, both via aquaculture or mariculture techniques.
    The ecological and socio-economic impacts of the ornamental trade 
industry for corals are numerous, and can include overharvesting, 
collateral damage to coral reef habitat, and potential introduction of 
exotic species (Rhyne et al., 2012). Wild collection of stony corals is 
usually highly destructive, resulting in removing and discarding large 
amounts of live coral that often go unsold for various reasons. 
Additionally, collection techniques can be physically damaging to reef 
habitat around live corals. In a recent, thorough review of ecological 
impacts and practices of the coral reef wildlife trade, Thornhill 
(2012) identifies and describes five overarching potential impacts: (1) 
Effects on target population such as over-exploitation and local 
population extirpations; (2) habitat impacts such as reduced coral 
cover, diversity, and rugosity; (3) effects on associated species such 
as decreased abundance, biomass, and diversity of reef fish, 
invertebrates, and other species due to loss or destruction of habitat; 
(4) ecosystem impacts such as increased degradation and erosion leading 
to reduced resilience; and (5) socio-economic impacts such as user 
group conflict between tourists, fishers, etc.
    Collection and trade of coral colonies can also increase the 
likelihood of the colonies being infected by disease, as a result of 
both the directed and incidental breakage of colonies, which are then 
more easily infected (Brainard et al., 2011). Further, destructive 
practices for collection of other coral reef wildlife, such as the use 
of cyanide for capturing reef fish, can also have deleterious effects 
on coral reef habitat in general. Currently, cyanide fishing is 
practiced in 15 countries, many of which are major marine wildlife 
trade exporters (Thornhill, 2012). There are likely many other examples 
of cumulative and interactive effects of collection and trade that pose 
a threat to corals. Given the paucity of data for the coral reef 
wildlife trade, it is difficult to accurately estimate mortality rates 
directly resulting from collection practices (Thornhill, 2012).
    The rapid increase of coral reef species entering markets in the 
United States and Europe and the sustainability of the aquarium trade 
in terms of driving collection of wild specimens have been of great 
concern to governments, scientists, conservationists, and conscientious 
aquarium hobbyists alike (Olivotto et al., 2011; Rhyne and Tlusty, 
2012). However, production of marine wildlife for home aquaria (i.e., 
the aquarium hobbyist trade) through captive culture is an increasingly 
growing sector of the ornamental trade industry. Recently, advances in 
both aquaculture and mariculture propagation techniques show promise in 
shifting the demand of the ornamental trade industry away from wild-
collected corals to corals reared via captive-culture techniques. Such 
techniques are possible since many corals, especially fast-growing 
branching corals, are capable of asexual reproduction via a process 
known as fragmentation or ``fragging'' (Brainard et al., 2011; Rhyne et 
al., 2012). According to CITES import and export reports, maricultured 
corals accounted for approximately 20 percent of total live trade in 
2010 (Wood et al. 2012), but other studies suggest that captive-
cultured corals account for only 2 percent of the live coral trade 
(Thornhill, 2012).
    Globally, there are approximately two million aquarium hobbyists 
involved in a complex trade network that sells an estimated 50 million 
corals every year to use (Rhyne et al., 2012). According to the Florida 
Department of Agriculture and Consumer Services, there are 87 certified 
aquaculture facilities listing corals as a product in Florida alone. 
The study hypothesized that a notable decline in U.S. imports of corals 
occurred after 2006 as a result of increased domestic coral production 
as well as the global economic downturn. Import reports do not account 
for this ``hidden'' domestic production, and statistical tracking of 
this type of coral production is lacking (Rhyne et al., 2012). In 
addition to increasing domestic production of corals, some major source 
countries such as Indonesia are increasing production via mariculture 
activities to reduce wild collection pressure on coral reefs, and 
supporting coral farming as a potential alternative to fishing for reef 
fish and collection of wild corals (Pomeroy et al., 2006). For example, 
according to 2009 U.S. import reports, 26 percent of Acropora species 
were identified under CITES codes which indicated that these colonies 
were produced via captive-

[[Page 53902]]

culture techniques (Rhyne et al., 2012). However, since CITES codes are 
self-determined by exporter countries, there may be some 
inconsistencies in how those codes are applied (Wood et al., 2012). As 
of 2008, there were 55 coral farms scattered throughout the different 
provinces of Indonesia (Timotius et al., 2009); however, this number 
may be increasing since Indonesia's government has mandated companies 
and traders involved in the coral trade to utilize captive culture 
techniques in hopes of eventually phasing out wild collection of 
corals.
    There are a number of challenges associated with developing 
aquaculture or mariculture operations for coral species, including 
technical capacity and know-how, high capital investments and operating 
costs, and high levels of production risk (Ferse et al., 2012; Pomeroy 
et al., 2006). Culturing corals has not been an easy task, 
predominantly due to the lack of knowledge regarding reproductive and 
larval biology for most traded species (Olivotto et al., 2011). 
Further, most mariculture operations tend to focus predominantly on 
fast-growing corals, while successful propagation techniques for the 
popular slow-growing, large-polyp species have not yet been developed 
(Wood et al., 2012). There is also the increasingly popular trend of 
using ocean-based coral nurseries for the purposes of propagating coral 
fragments to a suitable size and subsequently out-planting those coral 
fragments on degraded reefs to aid in reef restoration efforts. These 
types of activities are also considered in the Conservation Efforts 
section of the rule.
    The export of marine organisms for the ornamental trade industry is 
a global industry. As described in the proposed rule, it is estimated 
that 1.5 million live stony coral colonies are collected from at least 
45 countries each year, with an estimated 11 to 12 million coral pieces 
(i.e., fragments from larger colonies) traded every year (Brainard et 
al., 2011; Wabnitz, 2003). In addition to live stony corals, 
approximately 13 to 40 million reef fish, four million pounds of dead 
coral skeleton, and nine to 10 million other invertebrates are 
extracted from coral reef ecosystems across the world (Thornhill, 
2012). For corals, trade can be broken down into several categories, 
including: Coral rock (i.e., rock and substrate that may have live 
settled coral polyps among other marine organisms), live wild coral, 
live maricultured coral, and dead coral skeleton. Yet, numbers of 
corals traded in these categories are very difficult to accurately 
estimate for a variety of reasons. First, corals are colonial, vary in 
size, and can be fragmented into many smaller pieces. Additionally, 
reporting of trade volume is inconsistent and varies between reporting 
pieces and weight, and live rock and corals are often confused with 
each other and misreported (Thornhill, 2012). Currently, Indonesia is 
the primary source country of live corals; it exports approximately one 
million corals annually and represents an estimated 91 percent of the 
global supply market as of 2005 (Bruckner and Borneman, 2006; 
Thornhill, 2012; Timotius et al., 2009). Other major exporters of 
scleractinian corals include Fiji, Solomon Islands, Tonga, and 
Australia. The largest importers of coral reef wildlife include the 
United States, European Union, and Japan. The United States accounted 
for an average of 61 percent of global imports from 2000-2010 (Wood et 
al., 2012). Imports of live corals into the United States taken 
directly from coral reefs (not from aquacultured or maricultured 
sources) increased by 600 percent between 1988 and 2007, while the 
global imports of live coral increased by nearly 1,500 percent 
(Brainard et al., 2011; Thornhill, 2012; Tissot et al., 2010). Import 
and export data shows overall increasing trends for trade of live coral 
pieces between 2000-2009, with a slight dip in 2010 (Wood et al., 
2012). In addition, undocumented, illegal live coral trade is estimated 
to represent approximately 25 percent of the legal trade level, 
although these numbers are difficult to estimate considering the 
secretive nature of the illegal trade (Jones, 2008; Thornhill, 2012).
    The international coral trade was established by 1950 and was 
dominated by the Philippines until 1977 when a national ban on wild 
collection and export was introduced (Wood et al., 2012). It was then 
that Indonesia surpassed the Philippines to provide the majority of 
corals to the market. In the 1980s and 1990s, the international coral 
trade still focused on the trade of dead coral skeletons for home 
d[eacute]cor and curios. In recent years, the focus has shifted to live 
corals for the marine reef aquarium trade due to increased interest in 
home aquaria and advances in coral husbandry in North America and 
Europe, as well as the advent of modern air cargo methods (Rhyne et 
al., 2012; Thornhill, 2012; Wood et al., 2012). As stated previously, 
there is a complex global trade network of approximately two million 
aquarium hobbyists that sells upwards of 50 million coral reef animals 
every year (Rhyne et al., 2012). Collection of corals for display in 
public aquaria for educational purposes represents a small portion of 
the coral reef wildlife trade, and public aquaria likely produce as 
many corals as they consume by using captive-culture techniques 
(Thornhill, 2012).
    There has been some significant progress in captive culture of 
coral species using aquaculture and/or mariculture for the purposes of 
trade. Still, commercial-scale production of most species currently 
suffers several technical bottlenecks, including the long and often 
arduous supply chain from ocean to aquarium (e.g., capture, collection, 
handling, and transport), which often results in mortality ranging from 
a few percent up to 80 percent. For example, in an analysis of 
confiscated coral shipments, a majority of the corals were found in 
poor condition. On the way to their final destination, coral colonies 
may experience significant temperature drops in the shipping water, 
poor water quality, and physical damage from repeated handling of the 
shipping boxes and bags resulting in mortality of a large proportion of 
colonies through subsequent bacterial infections (Jones, 2008). These 
non-reported rates of biomass loss may significantly underestimate the 
ecological impacts of the trade as more corals are collected to make up 
the losses (Cohen et al., 2013; Thornhill, 2012). Distinguishing 
between specimens collected under regulated conditions from those 
collected using illegal or destructive fishing practices is very 
difficult (Cohen et al., 2013; Wabnitz, 2003).
    Traceability and tracking of cultured corals versus wild-collected 
corals is extremely difficult as there is no morphological or 
biological difference between them, making distinction almost 
impossible (Olivotto et al., 2011). For example, a coral can be broken 
into fragments and labeled as cultured, when in fact it was collected 
from the wild. There is some evidence to suggest that culture of live 
corals has the potential to affect trends in the trade industry by 
reducing wild collection and provide an economically and financially 
feasible alternative livelihood for local communities in the Indo-
Pacific. Even so, coral mariculture development in the Indo-Pacific is 
still in its infancy and requires a number of conditions to be met in 
order for these operations to be commercially profitable, sustainable, 
and traceable (Cohen et al., 2013; Pomeroy et al., 2006). It is also 
important to note that not all species lend themselves to culture. In 
fact, only a small number of coral genera have the ability to be 
commercially cultured (Rhyne et al., 2012). According to some sources, 
approximately 98 percent of

[[Page 53903]]

live corals in the ornamental trade are still collected from the wild, 
with only 2 percent originating from captive bred sources such as coral 
farms and nurseries (Ferse et al., 2012; Thornhill, 2012), but, 
according to a different analysis of import reports between 2000 and 
2010, captive cultured corals made up approximately 20 percent of total 
imports, and these originated almost entirely from Indonesia (Wood et 
al., 2012). Therefore, there are still significant data deficiencies 
and a large amount of uncertainty as to how much of an impact captive 
cultured corals are having on the ornamental trade.
    Significant supplemental information was received in public 
comments on the proposed rule or otherwise gathered on collection and 
trade of coral species. As previously described in the SRR and proposed 
rule, there are numerous ecological impacts from the physical process 
of removing corals and other wildlife from the reef. Trade practices 
that rely on the collection of wild individuals may damage or destroy 
adult and juvenile reef corals. Additionally, removal of reef fish and 
other organisms for trade purposes may also result in ecological 
impacts to reef ecosystems (Brainard et al., 2011). The ten most 
popular coral genera involved in the ornamental trade by volume are: 
Acropora (Indo-Pacific only), Euphyllia, Goniopora, Trachyphyllia, 
Plerogyra, Montipora, Heliofungia, Lobophyllia, Porites, and Turbinaria 
(Jones, 2008; Thornhill, 2012), all of which represent 31 of the coral 
species considered in this final rule. Acropora species are in the 
highest demand followed by the large polyp species such as Euphyllia 
(Jones, 2008). Culturing corals through aquaculture and/or mariculture 
techniques is becoming an increasingly popular tool to help move the 
aquarium trade away from collection of wild corals. Still, these 
techniques are fairly new and in need of many improvements before being 
considered a viable solution in shifting market demand from wild-
collected to captive cultured corals. As it currently stands, the 
amount of unreported, illegal, and unregulated collection, combined 
with the large amount of biomass loss along the supply chain raises 
serious questions as to the sustainability of the ornamental trade 
(Cohen et al., 2013). Overall, collection and trade of coral reef 
wildlife is considered to contribute to some individual species' 
extinction risk.
    In our previous analysis, collection and trade were generally 
considered to be a threat to coral reefs, as well as particular 
individual coral species, but extinction risk as a result of collection 
and trade activities for the 65 corals proposed for ESA listing was 
considered to be ``low'' (Brainard et al., 2011). After considering 
this supplemental information in addition to that which was available 
for the proposed rule, our conclusion regarding the threat of 
collection and trade remains unchanged from the proposed rule. 
Collection and trade is a low importance threat in assessing global 
extinction risk for the 65 corals in this final rule, and even less so 
for the seven Caribbean species due to undesirable appearance and 
growth characteristics for trade. The impact of collection and trade 
may be mediated by several factors and the extent to which the 
extinction risk of a coral species is impacted by collection and trade 
depends on its particular level of susceptibility, combined with its 
spatial and demographic characteristics in the context of worsening 
environmental conditions out to 2100, which is considered for each 
species in the Species Information and Determinations section. 
Information regarding the adequacy of regulations related to the marine 
ornamental trade such as CITES and other laws can be found in the Local 
Regulatory Mechanisms section of the Final Management Report (NMFS, 
2012b). Additionally, coral restoration projects using ocean-based, 
nursery-reared corals are also becoming increasingly popular as a 
complement to existing management tools. Information related to the 
roles that coral farms, coral nurseries, and aquaria (both public and 
private) play in coral reef conservation is discussed in the 
Conservation Efforts sub-section of the rule.

Inadequacy of Existing Regulatory Mechanisms (ESA Factor D)

    Regulatory mechanisms are considered under Factor D--Inadequacy of 
Existing Regulatory Mechanisms. As previously described in the proposed 
rule, we developed a Draft Management Report to assess the contribution 
of ``inadequacy of regulatory mechanisms'' to the extinction risk of 
corals. The Draft Management Report identified: (1) Existing regulatory 
mechanisms relevant to threats to the 82 candidate coral species; and 
(2) conservation efforts with regard to the status of the 82 candidate 
coral species. This Draft was peer reviewed and released with the SRR 
in April 2012, with a request for any information that we may have 
omitted. We incorporated all of the information we received into the 
Final Management Report, which formed the basis of our evaluation of 
this factor's effect on the extinction risk of the 82 candidate coral 
species in the proposed rule.
    The Final Management Report identified existing regulatory 
mechanisms that were relevant to the threats to coral species. It was 
organized in two sections: (1) Existing regulatory mechanisms that are 
relevant to addressing global-scale threats to addressing other threats 
to corals. The proposed rule summarized the information from that 
report as follows.
    Greenhouse gas emissions are regulated through multi-state 
agreements, at the international level, and through statutes and 
regulations, at the national, state, or provincial level. One of the 
key international agreements relevant to attempts to control GHG 
emissions, the Copenhagen Accord, was developed in 2009 by the 
Conference of Parties to the United Nations Framework Convention on 
Climate Change. The Copenhagen Accord identifies specific information 
provided by Parties on quantified economy-wide emissions targets for 
2020 and on nationally appropriate mitigation actions to the goal of 
capping increasing average global temperature at 2 [deg]C above pre-
industrial levels. Overall, the proposed rule concluded that existing 
regulatory mechanisms with the objective of reducing GHG emissions were 
inadequate to prevent the impacts to corals and coral reefs from ocean 
warming, ocean acidification, and other climate change-related threats. 
After an in-depth analysis of international agreements to curb GHG 
emissions and their respective progress, it appeared unlikely that 
Parties would be able to collectively achieve, in the near term, 
climate change avoidance goals outlined via international agreements. 
Additionally, none of the major global initiatives appeared to be 
ambitious enough, even if all terms were met, to reduce GHG emissions 
to the level necessary to minimize impacts to coral reefs and prevent 
what are predicted to be severe consequences for corals worldwide. The 
evidence suggested that existing regulatory mechanisms at the global 
scale in the form of international agreements to reduce GHG emissions 
were insufficient to prevent widespread impacts to corals.
    Existing regulatory mechanisms directly or indirectly addressing 
the localized threats identified in the proposed rule (i.e., those 
threats not related to GHGs and global climate change) are primarily 
national and local fisheries, coastal, and watershed management laws 
and regulations in the 84 countries within the collective ranges of the 
82 coral species. Because of the large number of threats, and the

[[Page 53904]]

immense number of regulatory mechanisms in the 84 countries, we 
concluded in the proposed rule that a regulation-by-regulation 
assessment of adequacy was not possible. Furthermore, with the 
exception of Acropora palmata and A. cervicornis in the Caribbean, 
there was not enough information available to determine the effects of 
specific regulatory mechanisms on individual coral species, given the 
lack of information on specific locations of individual species (the 
adequacy of existing local regulatory mechanisms relevant to threats 
impacting the Caribbean acroporids was evaluated in detail in those 
species' 2005 status review, and that information is incorporated into 
this rule's final findings for those species). However, general 
patterns included: (1) Fisheries management regimes regulate reef 
fishing in many parts of the collective ranges of the proposed coral 
species, albeit at varying levels of success; (2) laws addressing land-
based sources of pollution are less effective than those regulating 
fisheries; (3) coral reef and coastal marine protected areas have 
increased several-fold in the last decade, reducing some threats 
through regulation or banning of fishing, coastal development, and 
other activities contributing to localized threats; and (4) the most 
effective regulatory mechanisms address the threats other than climate 
change. We generally concluded that because the local threats have 
impacted and continue to impact corals across their ranges, 
collectively, the existing regulations were not preventing or 
controlling local threats. Further, there was insufficient information 
to determine if an individual species was impacted by inadequacy of 
individual existing regulations.
    We received public comments and supplemental information on the 
inadequacy of existing regulatory mechanisms. As a result, we 
incorporated any information we received into this final rule, which 
supplemented the basis for our final analysis and determination of the 
inadequacy of existing regulatory mechanisms in each species 
determination.
    Comments 31-33 provided supplemental information, which we 
incorporated into this final rule. Specifically, we received 
information on how local management actions potentially confer 
resilience benefits to coral reef ecosystems. The public comments and 
supplemental information on the inadequacy of existing regulatory 
mechanisms are discussed below in three sections: (1) Updates to 
adequacy of global regulatory mechanisms; (2) updates to adequacy of 
local regulatory mechanisms; and (3) local management as it applies to 
reef resilience.
    Since the release of the Final Management Report, there have been 
two additional conferences of the Parties to the United National 
Framework Convention on Climate Change. In 2012, the Parties met in 
Doha, Qatar, and they met again in Warsaw, Poland in 2013. The 
resulting decisions from both meetings were primarily to continue 
ongoing efforts to reach a new agreement for emissions reductions to be 
adopted at the 2015 meeting in Paris, and to have those implemented by 
2020. The new agreement would maintain the same overall goal as the 
Copenhagen Accord, to cap additional warming at 2 [deg]C. Within the 
United States, President Barack Obama released the President's Climate 
Action Plan in June 2013. The plan is three-pronged, including proposed 
actions for mitigation, adaptation, and international leadership. The 
actions listed for mitigation include completing carbon pollution 
standards for new and existing power plants, accelerating clean energy 
permitting, increasing funding for clean energy innovation and 
technology, increasing fuel economy standards, increasing energy 
efficiency in homes businesses and factories, and reducing other GHG 
emissions including hydrofluorocarbons and methane. The plan states 
that the United States is still committed to reducing GHG emissions 17 
percent below 2005 levels by 2020 if all other major economies agree to 
similar reductions. Additional efforts made domestically related to 
climate change are more focused on facilitating adaptation to the 
impending changes to the environment due to climate change in order to 
maintain the country's natural and economic resources, but do not 
directly address the emission of GHGs.
    As described in the proposed rule, existing regulatory mechanisms 
directly or indirectly addressing all of the localized threats 
identified in the SRR (i.e., those threats not related to GHGs and 
global climate change) are primarily national and local fisheries, 
coastal, and watershed management laws and regulations in the 84 
countries within the collective ranges of the 65 coral species. This 
final rule incorporates any information we received via public comment 
regarding recent local regulatory mechanisms or local regulatory 
mechanisms that were either previously mischaracterized or 
inadvertently omitted. This includes some additions of various local 
laws as well as supplemental information regarding regulations 
pertaining to collection and trade of coral species. In addition, to 
better capture the breadth and scope of existing regulatory mechanisms 
on a species-by-species basis, we evaluated the presence and scope of 
five different categories of regulatory mechanisms in each of the 84 
countries throughout the ranges of the 65 corals in this final rule. 
These categories of laws include: General protection of corals, reef 
fishing, marine protected areas, wild collection, and pollution.
    For each coral species, we considered the relevant national laws, 
regulations, and other similar mechanisms that may reduce any of the 
threats described in our threat analyses for all countries in which the 
coral species has confirmed records of occurrence. To find each country 
where our 65 coral species have confirmed occurrence we used Veron's 
updated report on the listed coral species and their occurrence in 
various ecoregions (Veron, 2014). In considering countries' regulatory 
mechanisms, we give strongest weight to statutes and their implementing 
regulations and to management direction that stems from those laws and 
regulations.
    In analyzing local regulatory mechanisms available for each coral 
species, five general categories emerged: General coral protection, 
coral collection control, fishing controls, pollution controls, and 
managed areas. General coral protection regulatory mechanisms include 
overarching environmental laws that may protect corals from damage, 
harm, and destruction, and specific coral reef management laws. In some 
instances, these general coral protection regulatory mechanisms are 
limited in scope because they apply only to certain areas or only 
regulate coral reef damage and do not prohibit it completely.
    Coral collection regulatory mechanisms include specific laws that 
prohibit the collection, harvest, and mining of corals. In some 
instances, these coral collection regulatory mechanisms are limited in 
scope because they apply only to certain areas or are regulated but not 
prohibited.
    Pollution control regulatory mechanisms include oil pollution laws, 
marine pollution laws, ship-based pollution laws, and coastal land use 
and development laws. In some instances, pollution regulatory 
mechanisms are limited in scope because they apply only to certain 
areas or to specific sources of pollution.
    Fishing regulatory mechanisms include fisheries regulations that 
pertain

[[Page 53905]]

to reefs or regulations that prohibit explosives, poisons and 
chemicals, electrocution, spearfishing, specific mesh sizes of nets, or 
other fishing gear. In some instances, fishing regulatory mechanisms 
are limited in scope because they apply only to certain areas, or not 
all reef-damaging fishing methods are prohibited, or reef-damaging 
fishing methods are regulated but not prohibited.
    Managed area regulatory mechanisms include the capacity to create 
national parks and reserves, sanctuaries, and marine protected areas. 
In some instances, managed area regulatory mechanisms are limited in 
scope, primarily because the managed area provides limited protection 
for coral reefs, only small percentages of the countries' coral reefs 
are protected within the managed areas, or the managed areas are not 
well administered.
    The management results for each species can be found in the 
Species-Specific Information and Determination section of this rule. It 
should be noted that while some of these regulatory mechanisms were 
categorized as ``limited in scope,'' it does not necessarily mean they 
are inadequate under ESA section 4(a)(1) Factor D.
    We received a significant amount of information regarding the role 
of local management actions in building resilience into reef 
ecosystems. This section describes the emerging body of literature 
regarding the concept of reef resilience, defined as an ecosystem's 
capacity to absorb recurrent shocks or disturbances and adapt to change 
without compromising its ecological function or structural integrity. 
Until recently, the main drivers of coral reef decline included 
overfishing of herbivorous fish and nutrient loading from agriculture 
and other land-based sources of pollution. These stressors caused 
widespread changes in reef ecosystems over the past couple of 
centuries, and ultimately led to ecological shifts from coral-dominated 
systems to systems overrun by fleshy algae. These localized 
disturbances are now being compounded by climate change related 
threats, including increasingly frequent coral bleaching events as a 
result of ocean warming.
    Many factors contribute to coral reef ecosystem resilience, 
including ecosystem condition, biological diversity, connectivity 
between areas, and local environmental conditions (Marshall and 
Schuttenberg, 2006; Obura, 2005). Implementing local actions that 
either protect or strengthen these resilience-conferring factors has 
the potential to help coral reef ecosystems survive predicted increases 
in the frequency, duration, and severity of mass coral bleaching events 
(Obura, 2005) and may help reduce the extinction risk of some 
individual coral species.
    In terms of local management actions, many acute disturbances such 
as coral bleaching are out of the direct control of reef managers and 
cannot be mitigated directly. Actions that can be taken to build reef 
resilience and enhance reef recovery include reducing physical 
disturbance and injury as a result of recreational activities, managing 
local watersheds and coastal areas to prevent sedimentation and 
nutrient run-off, and reducing fishing pressures on important 
herbivorous fish (Jackson et al., 2014; Kennedy et al., 2013; Marshall 
and Schuttenberg, 2006; Mumby and Steneck, 2011). For example, a recent 
study shows that eutrophication can increase thermal stress on inshore 
reef communities and management actions to reduce coastal 
eutrophication can improve the resistance and resilience of vulnerable 
coastal coral reefs to ocean warming (Fabricius et al., 2013). 
Additionally, herbivorous fish play a crucial role in the recovery of 
coral reefs after major disturbance events. Severe warming and 
increases in ocean acidification alone can reduce resilience of coral 
reef ecosystems, particularly if those systems are already subject to 
overfishing of the key functional groups of herbivorous reef fishes and 
nutrient loading (Anthony et al., 2011; Bellwood et al., 2004). 
Elevated populations of herbivores have the potential to confer 
resilience benefits by encouraging greater niche diversification and 
creating functional redundancy. For example, it has been demonstrated 
that two complementary herbivore species were more successful at 
controlling algal blooms than a single species on its own, and 
management of herbivorous fish can help in reef regeneration after 
episodes of bleaching or disease that are impossible to locally 
regulate (Bellwood et al., 2004; Burkepile and Hay, 2008; Roff and 
Mumby, 2012). Conversely, even unexploited populations of herbivorous 
fishes do not guarantee reef resilience; therefore, some reefs could 
lose resilience even under relatively low fishing pressure (Cheal et 
al., 2010). Therefore, the entire suite of local threats and 
disturbances should be minimized through local management actions to 
ensure that reef resilience and recovery are also maximized. 
Establishing MPA networks is generally accepted as one of the more 
common management tools to help reduce impacts to coral reefs and build 
resilience (Burke et al., 2011; Keller et al., 2009).
    In a 2013 global review of 10,280 MPAs, it was found that 
approximately 2.93 percent of the world's oceans have MPA coverage; 
however, coverage does not necessarily equate to protection. Marine 
protected areas have often failed to prevent ongoing local threats such 
as overfishing due to management and/or design failure, as well as lack 
of local support, poor compliance, and inadequate resources to promote 
educational awareness and enforcement (Hughes et al., 2007; Hughes et 
al., 2010; Spalding et al., 2013). A study by the World Resources 
Institute found that only 6 percent of the world's reefs occur in 
effectively managed MPAs (Burke et al., 2011). Further, scientists are 
just beginning to understand spatial patterns of coral responses to 
disturbance. Efforts to identify coral reef areas with the greatest 
resilience are crucial for siting MPAs. This information has the 
potential to assist in future MPA design and management so that 
resistant patches of coral reef can be protected to ensure continued 
connectivity and subsequent recovery of nearby reefs that are less 
resistant. These strategies of tailoring management efforts across the 
marine environment depending on various responses to disturbance are 
still in their infancy, but it may eventually prove essential in 
adaptive management of reef resources in the face of future climate 
change-related disturbances (Mumby and Steneck, 2011). For these 
reasons, while MPAs are an important tool in response to the global 
degradation of coral reefs, they should not be considered a panacea 
(Hughes et al., 2007).
    In general, recent evidence suggests that management of local scale 
disturbances is essential to maintaining an adequate coral population 
density for successful reproduction and maintenance of genetic 
diversity and is therefore crucial to maintaining complex, bio-diverse 
coral reef ecosystems, given the predicted widespread impacts of 
climate change related threats (e.g., Anthony et al., 2011). The 
presence of effective local laws and regulations has the potential to 
help reduce impacts to coral reefs from threats on an ecosystem level, 
potentially extending the timeframe at which individual coral species 
may be in danger of extinction by providing a protective temporal 
buffer (i.e., resiliency). Some evidence suggests that local management 
actions, particularly of fisheries (specifically, no-take marine 
reserves) and watersheds, can enhance the ability of species, 
communities, and ecosystems to tolerate climate change-

[[Page 53906]]

related stressors, and potentially delay reef loss by at least a decade 
under ``business-as-usual'' rises in GHG emissions (Keller et al., 
2008; Kennedy et al., 2013). In the Caribbean especially, local 
regulation of fisheries for herbivorous fish species (specifically 
parrotfish) is deemed one of the most important local actions to 
safeguard coral reefs in the face of looming climate change threats 
(Jackson et al., 2014). It also has been strongly suggested that local 
management be combined with a low-carbon economy to prevent further 
degradation of reef structures and associated ecosystems (Birkeland et 
al., 2013; Kennedy et al., 2013).
    After considering this supplemental information in addition to that 
which was available for the proposed rule, our conclusion regarding the 
inadequacy of regulatory mechanisms addressing global threats to corals 
from GHG emissions remains unchanged from the proposed rule. That is, 
without any substantive changes in emissions reduction pledges from any 
major economies and without any noteworthy additional efforts to 
actually reduce GHG emissions, the supplemental information considered 
in this final rule regarding regulatory mechanisms does not change the 
previous analysis. We reach the same conclusions regarding local 
regulatory mechanisms as described in the proposed rule, with the 
exceptions of Acropora palmata and A. cervicornis. For these species, 
we have incorporated into this final rule, the analysis of adequacy of 
regulatory mechanisms included in the 2005 status review and 2006 
listing of these species as threatened. Those documents concluded that 
existing regulatory mechanisms are inadequate to address local and 
global threats affecting these species, and as such are contributing to 
the threatened status of these species.
    Because the local threats have impacted and continue to impact 
corals across their ranges, we still generally conclude that, 
collectively, the existing regulations are not currently preventing or 
controlling local threats across the entire range of any of the 65 
species. We still do not have sufficient information to determine if an 
individual species' extinction risk is exacerbated by inadequacy of 
individual existing regulations. On the other hand, the best available 
information suggests that local management may confer resilience 
benefits for coral reefs on an ecosystem level, which could extend the 
timeframe at which individual coral species may be at risk of 
extinction by providing a protective temporal buffer in the face of 
climate change-related threats. That is, implementing effective local 
management actions may allow for coral to persist while awaiting 
significant global progress to curb GHGs. Overall, we maintain that in 
the absence of effective global regulatory mechanisms to reduce impacts 
from climate change to corals, the inadequacy of existing regulatory 
mechanisms at global and local scales poses an extinction risk threat 
to all of the corals that are vulnerable to climate-related threats.

Threats Evaluation Conclusion

    The above information on threats to reef-building corals leads to 
several important overall points that apply both currently and over the 
foreseeable future. First, the period of time over which individual 
threats and responses may be projected varies according to the nature 
of the threat and the type of information available about that threat 
and the species' likely response. The threats related to global climate 
change pose the greatest potential extinction risk to corals and have 
been evaluated with sufficient certainty out to the year 2100. Second, 
we expect an overall increase in threats, especially those related to 
global climate change as projected by RCP8.5 to 2100. Third, RCP8.5's 
projections of conditions on coral reefs within the ranges of the 
species covered by this rule over the foreseeable future are based on 
spatially-coarse analyses associated with high uncertainty, in 
particular at local spatial scales. Finally and most importantly, 
determining the effects of global threats on an individual coral 
species over the foreseeable future is complicated by the combination 
of: (1) Uncertainty associated with projected ocean warming and 
acidification threats; (2) regional and local variability in global 
threats; (3) large distributions and high habitat heterogeneity of the 
species in this final rule; and (4) limited species-specific 
information on responses to global threats.
    Thus, in our species determinations, we recognize that the best 
available information indicates the impacts of climate change will 
likely increase in the foreseeable future. However, there are 
limitations to using this global, coarse-scale information for 
determining vulnerability to extinction for individual coral species. 
Climate change projections over the foreseeable future are associated 
with three major sources of uncertainty; (1) The projected rate of 
increase for GHG concentrations; (2) strength of the climate's response 
to GHG concentrations; and (3) large natural variations. The recent 
warming slow-down is an example of a large natural variation that was 
not anticipated by previous models. Reports that discuss the future 
impacts of climate change on coral reefs indicate variability in both 
the models underlying these changes and the extent of potential impacts 
to the coral ecosystem. Recognizing uncertainty and spatial variability 
in climate change projections, and the spatial variability in 
environmental conditions on coral habitat, in our species 
determinations we emphasize the role that heterogeneous habitat and 
spatial and demographic traits play in evaluating extinction risk. We 
also consider in our determinations that each species in this final 
rule experiences a wide variety of conditions throughout its range that 
helps mitigate the impacts of global and local threats to some degree. 
Finally, we don't consider projections of impacts to coral reef 
ecosystems to definitively represent impacts to individual coral 
species, because coral reef communities typically consist of dozens to 
hundreds of reef-building coral species, each of which may respond 
differently to environmental and ecological changes. In addition, reef-
building corals are not limited to occupying only coral reefs.

Risk Analyses

    Many factors can contribute to an individual species' extinction 
risk. The process of extinction usually occurs in phases, first 
affecting individual populations or sub-populations, and then 
progressing to the species level. Extinction can occur as a result of 
stochastic processes that affect birth and death and mortality from 
catastrophic events. A species' biological traits can influence 
extinction risk both in terms of vulnerability to environmental 
perturbations and effects on population dynamics. Extinction risk is 
also influenced by depensatory effects, which are self-reinforcing 
processes (i.e., positive feedbacks) that accelerate species loss as 
its population density declines.
    The proposed rule described our framework for evaluating extinction 
risk and making listing determinations in the Risk Analyses section. 
There were multiple steps in our process of evaluating the listing 
status of each species. The initial step in developing the framework 
consisted of evaluating the ESA definitions of ``endangered'' and 
``threatened'' and how those definitions apply to corals. The 
application of those definitions was based on the background of the 
Context for Extinction Risk and General Threats sections of the 
proposed rule.
    We then considered the elements that contribute to the extinction 
risk of corals in the Risk Analyses section of the proposed rule. The 
following is a list

[[Page 53907]]

of the specific elements within their respective categories: (1) 
Vulnerability to threats, including each of the nine most important 
threats, based on a species' susceptibility and exposure to each of the 
threats; (2) demography, including abundance, trends in abundance, and 
relative recruitment rate; and (3) spatial structure, including overall 
distribution, which is a combination of geographic and depth 
distributions, and ocean basin. In order to evaluate the best available 
information for each of the 82 candidate corals and consider all 
elements in each of these categories, we developed a Determination Tool 
to organize and consistently interpret the information in the SRR, FMR, 
and SIR and apply it to the definitions of threatened, endangered, and 
not warranted species developed for corals, in a decision framework 
that we developed to specifically apply to corals.
    In the proposed rule, we linked the major elements of our Risk 
Analyses, vulnerability to threats, demography, and spatial structure, 
to the ESA listing categories. We described endangered species as 
having a current extinction risk; they are highly vulnerable to one or 
more of the high importance threats and have either already been 
seriously adversely affected by one of these threats, as evidenced by a 
declining trend and high susceptibility to that threat, or they lack a 
buffer to protect them from serious adverse effects from these threats 
in the future. We described threatened species as not currently being 
in danger of extinction, but are likely to become so within the 
foreseeable future. They are highly or moderately vulnerable to one or 
more of the high importance threats or highly vulnerable to one or more 
of the lower importance threats, but have either not yet exhibited 
effects in their populations or they have the buffering protection of 
more common abundance or wider overall distribution. We described not 
warranted species as not being in danger of extinction currently and 
not likely to become so within the foreseeable future because they 
have: Low vulnerability to the high importance threats, or low or 
moderate vulnerability to all the lower importance threats, and common 
abundance or wide overall distribution.
    The proposed rule described the basis for our determination of the 
foreseeable future for the purposes of projecting climate-related 
threats in the Threats Evaluation and Risk Analyses sections, and was 
supported by several other sections (e.g., Global Climate Change--
Overview). Consistent with our practice for all species listing 
determinations, we established that the appropriate period of time 
corresponding to the foreseeable future is a function of the particular 
type of threats, the life-history characteristics, and the specific 
habitat requirements for the coral species under consideration. The 
timeframe established for the foreseeable future considered the time 
necessary to provide for the conservation and recovery of each 
threatened species and the ecosystems upon which they depend. It was 
also a function of the reliability of available data regarding the 
identified threats and extends only as far as the data allow for making 
reasonable predictions about the species' response to those threats. We 
agreed with the BRT's assessment that the threats related to climate 
change had been sufficiently characterized and predicted through the 
end of this century. Therefore, in the proposed rule, we determined the 
year 2100 to be the appropriate outer limit of foreseeability as to 
climate change-related threats.
    In the proposed rule, we evaluated each species throughout its 
entire range, because no SPOIRs were identified, and that assessment 
has not changed in the final rule as described further below in the 
Statutory Standards sub-section. While we did receive additional 
qualitative information on the abundances and distributions of the 65 
proposed species, nothing in that data indicated that any portions of 
the range of any of the species warranted further evaluation under the 
applicable standards of the final SPOIR Policy, as discussed in the 
Statutory Standards sub-section below. The last step in developing the 
proposed listing determinations was to evaluate ``Conservation 
Efforts'' to determine if they would change the basis for listing a 
species by alleviating threats or recovering populations. We concluded 
that conservation efforts on global and local scales did not change the 
status determined using our decision framework for any of the 82 
candidate species.
    Comments 32-34 and 37-42 focused on four aspects of the listing 
determination process in the proposed rule: (1) The Determination Tool, 
(2) the foreseeable future, (3) the SPOIR analysis, and (4) 
conservation efforts. The comments we received identified deficiencies 
in the proposed rule's Determination Tool, leading to a change in our 
approach from a formulaic framework to describe extinction risk, to a 
non-formulaic framework to describe vulnerability to extinction. That 
is, the final determination framework integrates different types of 
information in a holistic manner that better represents all the 
available information, including complexity and uncertainty, than was 
possible using the linear Determination Tool in the proposed rule. In 
this section, we explain the final determination framework process that 
we used to determine each of the species' statuses, how it is different 
from the proposed rule, and how new and supplemental information was 
incorporated.
    In the proposed rule we described our determination approach in the 
Risk Analyses and Detailed Description of Determination Tool Elements 
sections, in which we discussed the elements that affect a coral's 
extinction risk. Below we describe how that determination approach has 
been adapted for this final rule and applied to the Statutory 
Standards, in light of and in response to public comments.

Final Determination Framework

    Overview of Key Changes Applied in Final Determinations. We 
received many comments questioning the accuracy of the methods used to 
analyze the available information to assess extinction risk and derive 
listing statuses for each of the proposed species, including how the 
Determination Tool was used. After considering these comments, and as 
discussed above, our findings in the proposed rule were influenced by 
how we believed coral species would react to environmental changes now 
and over the foreseeable future. Given the current effects and 
projections of climate change impacts to the marine environment into 
the foreseeable future and the information we had at the time of the 
proposed rule on coral response to existing and predicted environmental 
stressors, we determined that many of the coral species met the 
definition of ``endangered species'' or ``threatened species.'' In 
explaining how the Determination Tool assessed risk and derived listing 
statuses we concluded that, as some public comments suggested, the 
Determination Tool was too linear and deterministic. This led to 
listing determinations in the proposed rule that were based, in large 
part, on applying the endangered and threatened standard to relative 
characteristics instead of applying the endangered and threatened 
standard to each individual species independently to determine their 
listing status.
    In this rule, we have changed our determinations for many of the 
species for two general reasons: (1) Informed by public comments, we 
refined the way we apply the available information to determine 
vulnerability to extinction; and (2) we received via public comments, 
or gathered ourselves,

[[Page 53908]]

information that expanded our existing knowledge.
    We received and gathered specific information about spatial, 
demographic, and other characteristics of individual coral species, and 
the public comments provided general scientific criticism about how we 
weighed these factors. In the proposed rule, we gave greater 
consideration to susceptibility to threats but did not fully recognize 
the extent to which spatial, demographic, and other characteristics of 
corals can moderate vulnerability to extinction. After considering all 
of the available information and public comments, in this final rule we 
continue to recognize the threats that the species face, but we also 
place more emphasis on buffers against those threats and revisit the 
predicted population responses of individual species to the threats, 
giving full consideration to their current spatial, demographic, and 
other characteristics. For example, we took into account that many of 
the species, when viewed on their own rather than in relation to other 
coral species or vertebrate species, have more substantial absolute 
abundances than the prior methodology accounted for.
    We also took into account that in many instances coral species 
occupy a wide range of habitats, including areas that can act as 
refugia from warming, which moderate the predicted impacts across 
coarse-scale areas. As explained generally above, and in regard to 
individual species below, the species in this final rule will be 
negatively impacted by future conditions, but in light of our 
consideration of factors and characteristics discussed above, we find 
they are not currently in danger of extinction and do not meet the 
definition of endangered. We do, however, conclude that some species 
are likely to become in danger of extinction within the foreseeable 
future and thus meet the definition of threatened. We also find that 
listing is not warranted for some species that were previously proposed 
for listing.
    In this final rule, we acknowledge that there are no recipes or 
formulas for endangered, threatened, or not warranted coral species, 
especially given the variability in coral species' biology and ecology, 
and the variability in available information from species to species. 
Accordingly, the final framework allows for consideration of each 
coral's circumstances as a whole (simultaneously evaluating each 
species' demography, spatial characteristics, threat susceptibilities, 
and current and future environmental conditions independently of the 
other species), leading us to species-specific conclusions about 
vulnerability to extinction.
    The final determination framework used in this final rule is 
composed of seven elements. The first element is describing the 
statutory standards. The second, third, fourth, and fifth elements are 
identifying and analyzing all the appropriate species-specific and 
general characteristics that influence extinction risk for a coral 
species. The sixth element is relating a species' characteristics to a 
particular extinction risk at appropriate spatial and temporal scales. 
The seventh element is explicitly stating how each species' extinction 
risk meets the statutory listing definitions as applied to corals, 
resulting in an ultimate listing status. A final consideration in 
evaluating listing status is whether current or planned conservation 
efforts improve the overall status of any of the 65 species such that 
the additional protections of the ESA are not warranted.
    In moving to an integrated, non-formulaic framework, some of our 
key assumptions about vulnerability to extinction changed due to 
analyzing the different aspects of each species' characteristics 
independently (on an absolute scale), instead of being rated with the 
other proposed corals species (on a relative scale). We rely on the 
following guiding principles extracted from each of the sections in the 
first part of this rule, providing the context and background 
information for the species determinations, in order to determine each 
species' listing status:
     Clonal, colonial organisms, such as corals, are vastly 
different in their biology and ecology than many other species listed 
by NMFS under the Endangered Species Act.
     In our species determinations, we give appropriate 
consideration to the complex nature of coral biology and variability in 
responses to threats between individual coral colonies and even between 
different portions of the same colony.
     In our species determinations, absolute abundance and 
absolute distribution inform our evaluation of a species' current 
status and its capacity to respond to changing conditions over the 
foreseeable future.
     The concept of heterogeneous habitat influences extinction 
risk for all species in this final rule because each species 
experiences a wide variety of conditions throughout its range, which 
allows for variable responses to global and local threats.
     We recognize that the best available information indicates 
the impacts of climate change will likely increase in the foreseeable 
future. However, there are limitations to using this global, coarse-
scale information for determining vulnerability to extinction for 
individual coral species.
     In our species determinations, we don't consider 
projections of impacts to coral reef ecosystems to definitively 
represent impacts to individual coral species, because coral 
communities typically consist of dozens to hundreds of coral species, 
each of which may respond differently to environmental and ecological 
changes.
     Recognizing the uncertainty and spatial variability in 
climate change projections, and the spatial variability in 
environmental conditions on coral habitat, in our species 
determinations we emphasize the role that heterogeneous habitat and 
absolute demographic and spatial characteristics play in evaluating 
extinction risk.
    We have ordered the informational categories in the Species-
specific Information and Determinations sections below for clarity in 
describing the species-specific elements and their interaction in 
contributing to each species' vulnerability to extinction as follows: 
(1) Spatial Information--overall distribution and ocean basin, habitat; 
(2) Demographic Information--abundance, trends in abundance, relative 
recruitment rate; and (3) Susceptibility to threats based on a species' 
susceptibility to each of the nine threats. Further, when information 
is available that does not fall into one of the categories or elements 
identified above, but is relevant to extinction risk, we provide it 
under the Other Biological Information category. In each species 
determination, we refer back to the specific guiding principles that 
played a role in how we consider the species-specific information and 
the sections in which they are described in more detail.

Statutory Standards

    The definitions of endangered and threatened species under section 
3 of the ESA, wherein (1) an ``endangered species'' is defined as ``any 
species which is in danger of extinction throughout all or a 
significant portion of its range'', and (2) a ``threatened species'' is 
defined as ``any species which is likely to become an endangered 
species in the foreseeable future throughout all or a significant 
portion of its range,'' formed the basis of our determination 
framework. Considered at both the spatial and temporal scales 
applicable to each of those listing statuses, an endangered species 
currently faces an extinction risk throughout all or a significant 
portion of its range and a threatened species is likely to become 
endangered throughout

[[Page 53909]]

all or a significant portion of its range within the foreseeable 
future. In other words, the primary statutory difference between a 
threatened and endangered species is the timing of when a species may 
be in danger of extinction, either presently (endangered) or in the 
foreseeable future (threatened). Further, as discussed below, no 
significant portions of their ranges could be determined for any of our 
proposed species; thus, the only spatial scale we consider is each 
species' entire range.
    Court opinions produced in litigation challenging the listing of 
the polar bear as threatened provides a thorough discussion of the 
ESA's definitions and the Services' broad discretion to determine on a 
case-by-case basis whether a species is in danger of extinction (see, 
In Re Polar Bear Endangered Species Act Listing and Sec.  4(d) Rule 
Litigation, 794 F. Supp.2d 65 (D.D.C. 2011); aff'd, 709 F.3d 1 (D.C. 
Cir. 2013); 748 F. Supp.2d 19 (D.D.C. 2010)). The Court determined that 
the phrase ``in danger of extinction'' is ambiguous. The Court held 
that there is a temporal distinction between endangered and threatened 
species in terms of the proximity of the ``danger'' of extinction, 
noting that the definition of ``endangered species'' is phrased in the 
present tense, whereas a threatened species is ``likely to become'' so 
in the future. However, the Court also ruled that neither the ESA nor 
its legislative history compels the interpretation of ``endangered'' as 
a species being in ``imminent'' risk of extinction. Thus, in the 
context of the ESA, a key statutory difference between a threatened and 
endangered species is the timing of when a species may be in danger of 
extinction, either now (endangered) or in the foreseeable future 
(threatened). The Court ruled that although imminence of harm is 
clearly one factor that the Services weigh in their decision-making 
process, it is not necessarily a limiting factor, and that Congress did 
not intend to make any single factor controlling when drawing the 
distinction between endangered and threatened species. In many cases, 
the Services might appropriately find that the imminence of a 
particular threat is the dispositive factor that warrants listing a 
species as `threatened' rather than `endangered,' or vice versa. To be 
listed as endangered does not require that extinction be certain or 
probable, and that it is possible for a species validly listed as 
``endangered'' to actually persist indefinitely. Due to the ambiguous 
nature of the statutory terms, we have defined ``endangered'' and 
``threatened'' at the end of the Foreseeable Future sub-section below 
in the context of the particular species (corals) being considered for 
listing.
    Significant Portion of its Range (SPOIR). The ESA's definitions of 
``endangered species'' and ``threatened species'' refer to two spatial 
scales, providing that a species may be imperiled ``throughout all'' of 
or ``in a significant portion of'' its range. 16 U.S.C. 1532(6); (20). 
NMFS has interpreted the ``significant portion of its range'' language 
in a policy that has recently been finalized. See ``Final Policy on 
Interpretation of the Phrase `Significant Portion of its Range' in the 
Endangered Species Act's Definitions of `Endangered Species' and 
`Threatened Species' '' (79 FR 37578; July 1, 2014) (``Final Policy''). 
In developing our proposed rule, our analysis was informed by the Draft 
Policy that was published in December 2011 (76 FR 76987; December 9, 
2011). As we explained in the proposed rule, we were unable to identify 
any portions of the species' ranges that might require closer analysis 
as potential SPOIRs, due in large part to a lack of species-specific 
information regarding abundance, geographic distribution, diversity, 
and productivity (77 FR 73247).
    The Final Policy, which we must now apply, differs in two key 
respects from the Draft Policy. Neither changes the ultimate result in 
this case, which is that no SPOIRs can be identified. First, the Final 
Policy specifies that no portions of a species' range can be 
``significant,'' and thus no SPOIR analysis need be done, where the 
range-wide status analysis leads to a conclusion that listing the 
entire species as threatened or endangered is warranted. (Under the 
Draft Policy, even if a species were found to warrant listing as 
``threatened,'' the agency still needed to consider whether any 
portions of the range may be significant). Second, the final policy 
defines ``significant'' to include not only those portions where the 
individuals are so biologically significant that without them the 
entire species would meet the definition of ``endangered'' (the 
standard in the Draft Policy), but also those portions whose loss would 
render the species ``threatened.''
    In this case, our framework evaluates each species throughout its 
range to determine extinction risk. If a species is determined to be 
threatened or endangered based on the rangewide analysis, no further 
evaluation is warranted. However, if a species is found to be not 
warranted at the spatial scale of its entire range, we must consider if 
a SPOIR exists that may be both highly biologically important and at 
higher extinction risk, such that its loss would render the entire 
species endangered or threatened. An evaluation is required only where 
there is information to suggest that a particular portion of the range 
is likely to be both ``significant'' as defined in the policy and to 
qualify as endangered or threatened (79 FR 37586).
    As described in the proposed rule, the BRT did not identify any 
portions of the range for any of the 82 coral species as being 
potentially ``significant'' or at a higher extinction risk. Because 
there was a general lack of species-specific data regarding 
quantitative abundance, distribution, diversity, and productivity of 
coral species, we were not able to identify any portions of any of the 
species' ranges that could be considered unusually biologically 
significant. Further, we had no information to indicate that particular 
local threats were more severe in a particular portion of an individual 
species' range.
    No supplemental information was received in response to the 
proposed rule that provides support for identification of a SPOIR for 
any of the proposed species. While we did receive supplemental 
information on the qualitative abundances and distributions for some 
species, nothing in that data suggests that any particular portion of 
any proposed species range is unusually biologically significant. We do 
not have any information that would help elucidate whether any species 
is at higher exposure to threats in a particular area of its range 
(i.e., where threats may be so acute or concentrated that current 
conditions are likely to render the species there at significantly 
higher risk of extinction than the overall species). Thus, we did not 
identify any SPOIR for any species, and so our determination as to each 
species is based on the best available information about the species' 
status throughout its range.
    Foreseeable Future. The ``foreseeable future'' is integral to the 
definition of a threatened species. It is the timeframe over which we 
evaluate a species' extinction risk if it is not currently in danger of 
extinction. As described in the proposed rule, the identification of 
the foreseeable future is unique to every listing decision. It is based 
on the particular type of threats, the life-history characteristics, 
and the specific habitat requirements for the species under 
consideration.
    For this Final Rule, we clarify that the ``foreseeable future'' is 
that period of time over which we are able to make reliable projections 
about all of the significant threats affecting the species and the 
species' likely response to those

[[Page 53910]]

threats. Projections need not be ``certain'' to be reliable, so long as 
we are able to make predictions with a reasonable degree of confidence 
based on available information. In the proposed rule, we identified the 
year 2100 as marking the outer limit of the foreseeable future based 
upon the ability to make projections about the primary threats to 
corals--those stemming from global climate change--over that period (77 
FR 73226). However, in identifying 2100 as the limit of the foreseeable 
future for purpose of analyzing those threats, we did not intend to 
establish that year as the only relevant benchmark for analyzing all 
threats to the species or the species' response thereto.
    Because neither the ESA nor implementing regulations define 
``foreseeable future,'' the term is ambiguous, and Congress has left 
broad discretion to the Secretary to determine what period of time is 
reasonable for each species. This does not require identifying a 
specific year or period of time to frame our analysis, particularly 
where there is inadequate specific data to do so. See ``Memorandum 
Opinion: The Meaning of `Foreseeable Future' in Section 3(20) of the 
Endangered Species Act'' (M-37021, Department of the Interior Office of 
the Solicitor, January 16, 2009). The appropriate timescales for 
analyzing various threats will vary with the data available about each 
threat. In making our final listing determinations we must synthesize 
all available information and forecast the species' status into the 
future only as far as we reliably are able based on the best available 
scientific and commercial information and best professional judgment.
    In the case of corals, we can make reasonable assessments as to the 
most significant environmental factors facing the coral species between 
now and 2100. We have explained that this time period, which is 
consistently used by most current global models and the IPCC reports, 
allows for reliable and reasonable projections about climate change-
related threats. As described in the Threats Evaluation--Foreseeable 
Future and Global Climate Change Overview sections above, 2100 was 
selected as the limit of foreseeability for climate change-related 
threats based on AR4's and AR5 WGI's use of 2100 as the end-point for 
most of its global climate change models (IPCC, 2013). Public comments 
asserted that the models used in climate predictions are too uncertain 
to reliably predict climate conditions out to 2100. However, as we have 
explained in our response to Comment 38 and elsewhere in this final 
rule, supplemental information supports, and we reaffirm our choice of, 
identifying 2100 as the timeframe over which we can make reliable 
predictions about climate change-related threats.
    However, global climate change is not the only relevant threat to 
the species, and the range of available data differs as to these other 
threats (such as predation, sedimentation, etc.). Further, in reaching 
our conclusions and ultimate listing determinations, we need to assess 
how the species will react to the various stressors identified in this 
rule. For example, to the extent it was available, we considered a 
significant amount of information on the current spatial and 
demographic features of the species, based on various types of 
information which support varying degrees of projection into the 
future. Thus, while the year 2100 is a reliable end-point for 
projecting climate change-related threats, it is not valid across the 
range of threats for the species and should not be misunderstood as 
driving our forecasts of the species' statuses.
    For all of these species, we concluded based on the best available 
scientific and commercial information that their spatial, demographic, 
or other characteristics buffer them against current endangerment of 
extinction. However, over the foreseeable future, the ability of 
spatial and demographic traits to provide a buffer against the danger 
of extinction is expected to diminish as colonies within particular 
areas are impacted due to climate change and other negative stressors. 
We considered, at a species level, whether these predicted conditions 
may cause the species to become in danger of extinction within the 
foreseeable future. However, there are varying degrees of certainty 
about the responses of corals to stressors. We can be confident that 
certain mitigating elements of the life history for some of these 
species will not change, such as their ability to reproduce asexually 
or the ability to persist in a range of depths. But we are less 
confident in other aspects, such as precisely where and when local 
extirpations may occur.
    For this final rule, then, we make clear that our listing 
determinations are reached on the totality of the best available 
information about the threats to the species and the species' likely 
response to them over time. Our determinations reflect our 
consideration of that information, as well as application of our 
professional judgment regarding how far into the future we can reliably 
project either the underlying threats or the species' response. 
However, in light of the number of variables pertaining to the 
stressors and buffering traits among the corals species evaluated, and 
the limited availability and incomplete nature of quantitative data on 
these species, a quantitative assessment of these projections is not 
possible. Therefore our assessment of the foreseeable future is 
necessarily qualitative. Given the biological traits and life history 
strategies of the corals evaluated in this rule, including their 
relatively long life-spans, the period of time over which we are able 
to make reliable projections is the next several decades. This general 
timeframe thus frames our listing determinations. Although we recognize 
that climate related threats will persist beyond this horizon, we find 
it both infeasible on the information available and unnecessary to 
attempt to identify the foreseeable future across the full range of 
threats to the species and the species' response with more precision.
    In the proposed rule, we considered how the temporal scales were 
appropriately factored into our evaluations of whether a species was in 
danger of extinction now, likely to become in danger of extinction in 
the foreseeable future, or not warranted for listing. For example, two 
major factors determining the immediacy of the danger of extinction for 
corals are the relatively high degree of certainty of impacts from high 
importance threats and a species' current or future capacity to resist 
adverse effects. Under the proposed rule's Determination Tool approach, 
endangered species were species with a current high extinction risk; 
they were highly vulnerable to one or more of the high importance 
threats and had either already been seriously adversely affected by one 
of these threats, as evidenced by a declining trend, and high 
susceptibility to that threat, or they lacked a buffer to protect them 
from serious adverse effects from these threats in the future. While a 
threatened species under the proposed rule might be impacted by the 
same threats as an endangered species, it was less exposed or less 
susceptible, providing greater buffering capacity to those same threats 
when compared to an endangered species.
    In response to public comments critical of our equating species' 
listing statuses with outcomes of the determination tool, here we more 
fully explain the biological characteristics and distinctions between 
endangered and threatened corals, and corals not warranting listing 
under the ESA. Under the final rule's determination framework, an 
endangered species is at such risk of extinction, that it is currently 
``in danger'' of extinction throughout its range. As such, an 
endangered coral species is of such low

[[Page 53911]]

abundance or is so spatially fragmented that the species is currently 
in danger of extinction. Several processes may contribute to the danger 
of extinction (e.g., depensatory process, catastrophic events). 
Depensatory processes include reproductive failure from low density of 
reproductive individuals and genetic processes such as inbreeding. A 
coral species with these characteristics would be vulnerable to 
background environmental variation if a large proportion of the 
existing population were concentrated in an area that experienced an 
environmental anomaly leading to high mortality. Similarly, an 
endangered coral species could be of such low abundance that one 
catastrophic event or a series of severe, sudden, and deleterious 
environmental events could cause mortality of a large enough proportion 
of the existing population that the remaining population would be 
unable to reproduce and/or recover. A coral species that meets the 
endangered standard is not necessarily characterized by a single factor 
(e.g., abundance number, density, spatial distribution, or trend value) 
but could also be characterized by combinations of factors encompassing 
multiple life history characteristics and other important ecological 
features, as described above. Different combinations of such factors 
may result in endangered status from species to species.
    Under the final rule's determination framework, a threatened coral 
species also is at a risk of extinction due to its spatial and 
demographic characteristics and threat susceptibilities; however those 
traits still provide sufficient buffering capacity against being in 
danger of extinction currently. In other words, the species has an 
abundance and distribution sufficient for it to be not currently of 
such low abundance or so spatially fragmented to be in danger of 
extinction, but is likely to become so within the foreseeable future 
throughout it range. Similar to an endangered species, a coral species 
that meets the threatened standard is not necessarily characterized by 
a single factor (e.g., abundance number, density, spatial distribution, 
or trend value) but could also be characterized by combinations of 
factors encompassing multiple life history characteristics and other 
important ecological features, as described above. Different 
combinations of such factors may result in threatened status from 
species to species.
    Thus, there is a temporal distinction between endangered and 
threatened species in terms of the proximity of the danger of 
extinction based on the sufficiency of characteristics to provide 
buffering capacity against threats that cause elevated extinction risk. 
It is worth noting that this temporal distinction is broad, and a 
threatened species could likely become an endangered species anytime 
within the foreseeable future.
    Under the final rule's determination framework, a coral species 
that is not warranted for listing has spatial and demographic traits 
and threat susceptibilities that, when considered in combination, 
provide sufficient buffering capacity against being in danger of 
extinction within the foreseeable future throughout its range. In other 
words, it has sufficient abundance and distribution, when considering 
the species' threat susceptibilities and future projections of threats, 
it is not likely to become of such low abundance or so spatially 
fragmented to be in danger of extinction within the foreseeable future 
throughout its range. A not warranted species also may not be 
susceptible to the threats at a sufficient level to cause any major 
change in the species abundance.
    In summary, the basic structure of our final determination 
framework is formed by the relevant spatial and temporal scales over 
which each coral species' extinction risk is evaluated. An endangered 
coral species is currently in danger of extinction throughout its 
entire range. A threatened species is likely to become endangered 
throughout its entire range within the foreseeable future.

Spatial Structure

    We consider spatial elements that increase a species' risk of 
extinction, alone or in combination with other threats, under ESA 
Factor E--other natural or manmade factors affecting the continued 
existence of the species. Spatial structure is important at a variety 
of scales. At small spatial scales within a single population, issues 
of gamete density and other Allee effects can have significant impacts 
on population persistence. At large spatial scales, geographic 
distribution can buffer a population or a species from environmental 
fluctuations or catastrophic events by ``spreading the risk'' among 
multiple populations. We explicitly described how exposure to 
individual threats varies at different spatial scales in the Threats 
Evaluation section above. Generally, having a larger geographic or 
depth distribution provides more potential area to occupy. However, if 
populations become too isolated gene flow and larval connectivity may 
be reduced, making the species less likely to recover from mortality 
events. Thus, a robust spatial structure includes larger geographic 
distributions with adequate connectivity to maintain proximity of 
populations and individuals within the range. We consider geographic 
distribution and depth distribution (and connectivity, when we have 
that information) in describing the overall distribution for each 
species.
    We also consider the ocean basin in which a species exists. As 
described in the Corals and Coral Reefs--Inter-basin Comparisons, the 
Indo-Pacific occupies at least 60 million square km of water (more than 
ten times larger than the Caribbean), and includes 50,000 islands and 
over 40,000 km of continental coastline, spanning approximately 180 
degrees of longitude and 60 degrees of latitude. Thus, occupying only a 
small portion of the Indo-Pacific basin can still be a geographically 
large distribution for an individual coral species. In contrast, the 
Caribbean basin is relatively geographically small and partially 
enclosed, but biologically well-connected. The Caribbean also has 
relatively high human population densities with a long history of 
adversely affecting coral reef systems across the basin. In the 
proposed rule we determined that if a species is restricted to the 
Caribbean, its overall range was considered narrow and its extinction 
risk was significantly increased, which greatly contributed to an 
endangered or threatened determination. Comment 40 criticizes our 
characterization of the Caribbean in this manner, stating that the 
BRT's determination that the entire Caribbean is sufficiently limited 
in geographic scale to be a factor that increases the extinction risk 
of all corals in the Caribbean is at odds with genetic data. The 
commenter provided references to support the conclusion that, while it 
is clear that regional-scale processes such as bleaching and disease 
are acting on all these reefs in the Caribbean basin simultaneously, 
all reefs should not be presumed to respond the same to these 
disturbances. Upon consideration of the comment and the fact that the 
Determination Tool ratings regarding basin occupancy were an 
inadvertent function of comparing the Caribbean basin to Indo-Pacific 
basin (i.e., the automatic increase in extinction risk for species 
occurring in the smaller, more disturbed Caribbean was only relative in 
comparison to species occurring in the larger, less disturbed Indo-
Pacific) we re-evaluated our characterization of the Caribbean. We now 
consider the absolute (non-relative) size of the basin and the amount 
of heterogeneity in the system; therefore, we no longer

[[Page 53912]]

conclude that presence within the Caribbean basin automatically 
increases extinction risk (because many of the Caribbean coral species 
occupy a large portion of habitat compared to the total habitat 
available to them and the heterogeneous nature of that habitat). In 
general, we still consider distribution in the Caribbean to be 
problematic, but will now consider the influence of a Caribbean 
distribution on extinction risk on a species-by-species basis. For 
example, if a species has a Caribbean-wide geographic distribution and 
large depth distribution, and isn't susceptible to or exposed to 
threats now or through the foreseeable future, then a Caribbean basin 
distribution alone doesn't automatically increase the species' 
extinction risk. In the Species-specific Information and Determinations 
section of this final rule, we describe the extent to which an 
individual species' extinction risk is influenced by its specific 
geographic, depth, and habitat distributions within each basin.

Demography

    Demographic elements that cause a species to be at heightened risk 
of extinction, alone or in combination with threats under other listing 
factors, are considered under ESA Factor E--other natural or manmade 
factors affecting the continued existence of the species. In the 
proposed rule, we used species-specific qualitative abundance 
estimates, coded as ``common,'' ``uncommon,'' or ``rare'' for the 
candidate species because it was the only abundance metric that was 
available for all of the 82 candidate species. As mentioned above in 
the Distribution and Abundance of Reef-building Corals sub-section, 
these qualitative estimates are the subjective opinion of particular 
authors on their particular survey data and are meant to indicate 
relative abundance between the categories. That is, a rare species has 
fewer individuals as compared to an uncommon one, and an uncommon 
species has fewer individuals than a common one. These estimates are 
also meant to describe an author's opinion of the qualitative abundance 
of the species throughout its range, but not an estimate of the 
abundance at an individual location. In general, ``rare'' or 
``uncommon'' species are more vulnerable than ``common'' ones, although 
some species are naturally rare and have likely persisted in that rare 
state for tens of thousands of years or longer. However, naturally rare 
species can be at greater risk of extinction than naturally more common 
species when confronted with global threats to which they are 
vulnerable. In our final determination framework, rarity or 
uncommonness may increase extinction risk, but alone it does not 
automatically contribute to a finding of an endangered or threatened 
status.
    Trends in abundance directly demonstrate how a particular species 
responds under current or recent-past conditions. Generally, a 
continuing downward trend likely increases extinction risk, while 
stabilization or a continuing upward trend likely decreases extinction 
risk. Trend data for the 65 species are scarce, but we describe the 
extent to which an individual species' extinction risk is influenced by 
its trend data when the information is available.
    Productivity is another important indicator of extinction risk. 
Productivity is defined here as the tendency of the population to 
increase in abundance and is often expressed as ``recruits per 
spawner,'' although the term ``recruit'' can be difficult to apply in 
the case of corals, which reproduce both sexually and asexually (see 
Individual Delineation sub-section). Some of the proposed coral species 
are long-lived, with low or episodic productivity, making them 
vulnerable to trends of increased mortality or catastrophic mortality 
events. Overall, recruitment rate estimates for the proposed species 
are scarce, but in cases where estimates were available analysis of how 
that species' extinction risk is influenced by its relative recruitment 
rate is considered in the Species-specific Information and 
Determinations section below.

Susceptibility to Threats

    Susceptibility of a coral species to a threat is primarily a 
function of biological processes and characteristics, and can vary 
greatly between and within taxa. Susceptibility of a species to a 
threat depends on the combination of: (1) Direct effects of the threat 
on the species; and (2) the cumulative and interactive (synergistic or 
antagonistic) effects of the threat with the effects of other threats 
on the species. In the proposed rule, we considered how the cumulative 
or interactive effects altered the rating assigned to a threat 
susceptibility in isolation. However, upon further consideration, we 
need to evaluate the extent to which one threat influences the 
susceptibility of an individual species to another threat with more 
species-specific information, in connection with all the other elements 
that influence a species' extinction risk. Generally, cumulative and 
interactive processes are complex and uncertain and existing 
information about threats interactions is only based on a few studies 
on a few species. Where possible, when we have species-specific 
cumulative or interactive effects information, we have applied this 
information to that particular species' susceptibilities in a more 
integrated manner. Species-specific threat susceptibilities are 
described in the Species-specific Information and Determinations 
section.
    The three most important threats that contribute to the proposed 
coral species' extinction risk are ocean warming, disease, and ocean 
acidification. We considered these threats to be the most significant 
threats posing extinction risk to the proposed coral species currently 
and out to the year 2100. Threats of lower importance (trophic effects 
of reef fishing, sedimentation, nutrients, sea-level rise, predation, 
and collection and trade) also contributed to our findings on 
extinction risk, but to a lesser extent.

Current and Future Environmental Conditions

    The general information described in the preceding sections of this 
final rule illustrates that the most important threats are currently 
increasing and likely to increase further in the foreseeable future 
(Threats Evaluation), but that the impacts from these threats currently 
and in the foreseeable future are difficult to interpret and do not 
necessarily correlate to an increased vulnerability to extinction due 
to the biological and physical complexity of corals and their habitat 
(Corals and Corals Reefs, Threats Evaluation).
    The information on corals, coral reefs, coral habitat, and threats 
to reef-building corals in a changing climate leads to several 
important points that apply both currently and over the foreseeable 
future. First, the foreseeable future for purposes of our ultimate 
listing determinations is described qualitatively and encompasses the 
next several decades. For purposes of analyzing the specific threats 
related to climate change, we have identified the foreseeable time 
period over which we can make reliable projections to extend over the 
period from now to the year 2100. There is increased uncertainty over 
that time period as conditions that are analyzed closer to the year 
2100 become less foreseeable. That is, the general trend in conditions 
during the period of time from now to 2100 is reasonably foreseeable as 
a whole, but conditions become more difficult to accurately predict 
through time. Second, there is an overall increasing trend of threat 
severity, especially for threats related to global climate change as 
projected by RCP8.5 to 2100. Third, while some models suggest 
disastrous

[[Page 53913]]

effects of RCP8.5 on coral reefs by 2100, such projections are based on 
spatially coarse analyses associated with high uncertainty, especially 
at local spatial scales. In sum, determining the effects of global 
threats on an individual coral species over the foreseeable future is 
complicated by the combination of: (1) Uncertainty associated with 
projected ocean warming and acidification threats; (2) regional and 
local variability in global threats; (3) large distributions and high 
habitat heterogeneity of the species in this final rule; and (4) 
limited species-specific information on responses to global threats.

Vulnerability to Extinction

    The vulnerability of a species to extinction is a complex function 
of physiology, life history, morphology, spatial distribution, and 
interaction with threats (the biological context). The biological 
context for a species' vulnerability to threats dictates the ecological 
interactions that ultimately determine how a species responds to 
threats, such as competition and predation (the ecological context). 
For example, a species that suffers high mortality from a bleaching 
event also may be able to recover quickly because its high dispersal 
and skeletal growth enable efficient recolonization and strong 
competition. Thus, the initial response to threats does not necessarily 
mean the species is vulnerable.
    Vulnerability of a coral species to extinction also depends on the 
proportion of colonies that are exposed to threats and their different 
responses to those threats. In the proposed rule there was little 
variation between species for exposure to a given threat in the 
assigned ratings (e.g., exposure to ocean warming was rated the same 
for all 82 species, which should not automatically be the case because 
for species that have drastically different distributions and 
abundances). For this final rule, a coral species' vulnerability to 
extinction is now evaluated to be holistically influenced by its 
demographic and spatial characteristics, threat susceptibilities, and 
current and future environmental conditions. We believe this more 
complete and integrated treatment of the factors that influence a 
coral's vulnerability to extinction will lead to a more accurate 
characterization of whether or not a species currently faces an 
extinction risk throughout its entire range.

Species Status

    After analyzing all of the relevant species-specific demographic 
and spatial characteristics, threat susceptibilities, and general 
information on current and future environmental conditions, we relate 
those characteristics to the particular species' status. This is the 
key component of the determination that explains how certain species 
characteristics translate to a particular extinction risk at 
appropriate spatial and temporal scales. These determinations are 
heavily influenced by the quantity and quality of species-specific 
information, especially the species' demographic and distribution 
characteristics. We received many public comments regarding the lack of 
quantity and quality of information available for each of the species; 
those commenters asserted that our species determinations were 
therefore unfounded. By specifically considering all the currently 
available species-specific information (both information that we used 
in the proposed rule and the considerable amount of information that 
has become available since the proposed rule), we are able to produce 
more robust evaluations of the information and species determinations. 
Recognizing the uncertainty and spatial variability of climate change 
projections and the limited species-specific information on how species 
in this final rule respond to climate change, we emphasize a species' 
demographic and spatial characteristics in how its vulnerability to 
extinction is affected now and through the foreseeable future.
    In finalizing a species determination we translate the species' 
status directly into a listing category using the statutory standards. 
In the proposed rule, we satisfied this step by using an organizational 
process called the outcome key, based on ratings in the Determination 
Tool. The key was intended to identify the general species 
characteristics and combinations that equate to a particular listing 
status. However, the outcome key in the proposed rule was too 
formulaic, and did not explain our comprehensive consideration of the 
species characteristics that influenced their listing status, and was 
also based on relative ratings from the Determination Tool. Therefore, 
the presentation of our final determination framework is more clearly 
articulated in this final rule by explicitly describing the 
considerations for each the 65 species in narrative format and how they 
relate to the statutory standards
    In summary, the determination framework used in this final rule is 
composed of seven elements. The first element is describing the 
statutory standards. The second, third, fourth, and fifth elements are 
identifying and analyzing all the appropriate species-specific and 
general characteristics that influence extinction risk for a coral 
species. The sixth element is relating a species' characteristics to a 
particular extinction risk at appropriate spatial and temporal scales. 
The seventh element is explicitly stating how each species' extinction 
risk meets the statutory listing definitions as applied to corals, 
resulting in an ultimate listing status. A final consideration in 
evaluating listing status is whether current or planned conservation 
efforts improve the overall status of any of the 65 species such that 
the additional protections of the ESA are not warranted. We explicitly 
apply the determination framework to each species in our narrative 
evaluations. This approach provides consistency across all of the 65 
final listing determinations, but also produces individual 
determinations that are independent of the other 65 coral species.

Conservation Efforts

    The effect conservation efforts have on an individual species' 
listing status is the last consideration in making a final 
determination. Because many conservation efforts are not species-
specific, we provide our analysis of the effectiveness of conservation 
efforts for corals generally prior to making individual species 
determinations. Our conclusions regarding conservation efforts in this 
section apply to all of the proposed species. However, in some cases, 
we are able to identify species-specific conservation efforts and 
therefore evaluate them separately in the Species-specific Information 
and Determinations section.
    Section 4(b)(1)(A) of the ESA requires the Secretary, when making a 
listing determination for a species, to take into account those 
efforts, if any, being made by any State or foreign nation to protect 
the species. In evaluating the efficacy of protective efforts, we rely 
on the Services' joint ``Policy for Evaluation of Conservation Efforts 
When Making Listing Decisions'' (``PECE;'' 68 FR 15100; March 28, 
2003). The PECE requires us to consider whether any conservation 
efforts recently adopted or implemented, but not yet proven to be 
successful, will result in improving the species' status to the point 
at which listing is not warranted, or contribute to a threatened rather 
than endangered status.
    For the proposed rule, we developed a Management Report that 
identified existing conservation efforts relevant to both global and 
local threats to corals. A draft of this report was peer reviewed and 
made available to the public with the SRR in April 2012. At that time, 
we

[[Page 53914]]

requested any new or inadvertently overlooked existing information. The 
information that we received was incorporated into the Final Management 
Report (NMFS, 2012b), which formed the basis of our initial PECE 
evaluation. The information, analysis, and conclusions regarding 
conservation efforts in the proposed rule and supporting documents 
apply to this final rule, unless otherwise noted below.
    Comments 30-32 focus on our consideration of conservation efforts 
in the proposed rule. In response to public comments on the proposed 
rule, we incorporated into our analyses in the final rule relevant 
information on conservation efforts that are new or that may have been 
inadvertently omitted or mischaracterized. Thus, this final rule 
incorporates information we received as a result of the public comment 
period, identifies existing conservation efforts that are relevant to 
the threats to the 65 coral species in this final rule, both for 
global-scale threats to corals linked to GHG emissions and other 
threats to corals. In particular, we received supplemental information 
regarding coral reef restoration efforts in Florida and the wider-
Caribbean. We also received supplemental information regarding efforts 
to utilize captive-culture techniques to supplement the coral reef 
wildlife trade industry and reduce collection pressure on wild coral 
species. Specifically, we received information regarding Indonesia's 
mariculture operations as well as efforts in the United States to 
commercially and recreationally farm corals. This information on coral 
reef restoration, captive culture efforts for trade purposes, and local 
conservation efforts as it applies to reef resilience is described 
further below.
    We received some supplemental information regarding the ongoing 
coral reef restoration efforts being made in South Florida as well as 
the wider-Caribbean, predominantly for staghorn and elkhorn corals 
(Acropora cervicornis and A. palmata, respectively). We briefly 
mentioned active coral restoration in the proposed rule as an important 
conservation action for corals, but did not describe these efforts in 
great detail. Coral reef restoration efforts encompass a variety of 
activities, and they are increasingly utilized to enhance, restore, and 
recover coral reef ecosystems and species (Bowden-Kerby et al., 2005; 
Bruckner and Bruckner, 2001; Lirman et al., 2010b). These activities 
may include post-ship grounding ``triage'' (e.g., stabilizing substrate 
and salvaging corals and sponges), active predator and algae removal, 
larval seeding, and active restoration via coral propagation and 
outplanting activities. As a result of the 2009 American Recovery and 
Reinvestment Act, Federal funding through NOAA enabled a network of 
coral nurseries to expand throughout south Florida and the U.S. Virgin 
Islands to help recover threatened staghorn and elkhorn corals. These 
types of in-water coral nurseries have proven successful for 
propagating corals and serving as genetic repositories to help 
replenish and restore denuded reefs (Schopmeyer et al., 2012; Young et 
al., 2012). In 2012 alone, it was estimated these nurseries housed 
30,000 corals, with more than 6,000 corals outplanted to surrounding 
reefs (The Nature Conservancy, 2012). Further, successful spawning of 
these outplanted corals has been reported on several occasions since 
the first event occurred in 2009 (Coral Restoration Foundation, 2013). 
Still, it should be emphasized that coral reef restoration should not 
be expected to recover entire reef tracts or species; rather, coral 
reef restoration can serve as a complementary tool to other management 
strategies such as fisheries management, coastal zone and watershed 
management, marine protected areas, and others. In a comprehensive 
review of restoration activities conducted in Florida and the wider-
Caribbean, Young et al. (2012) found that most practitioners 
recommended that active restoration activities always be conducted in 
conjunction with robust local and regional management strategies to 
minimize the impacts of global and local threats. This is because coral 
reef restoration efforts can prove futile if the initial elements of 
degradation have not been mitigated (Jaap, 2000; Precht and Aronson, 
2006; Young et al., 2012).
    As described above in the Threats Evaluation--Collection and Trade 
section of this rule, we received a significant amount of information 
regarding the potential conservation benefits of increasing 
international and domestic commercial and recreational production of 
corals via significant advances in captive-culture techniques (i.e., 
mariculture and aquaculture). Specifically, we received supplemental 
information regarding the mariculture efforts conducted in Indonesia to 
reduce the amount of corals collected in the wild, thereby potentially 
reducing the threat of the marine ornamental trade industry on corals 
and coral reefs. As the largest exporter of corals in the world, 
shifting from wild-collected corals to captive cultured corals is an 
important conservation effort for preserving the integrity of wild 
reefs and coral species in Indonesia. However, there are still many 
challenges and obstacles related to captive culture of corals that are 
detailed in the Threats Evaluation, Trade and Collection section above. 
Any relevant information regarding this topic has also been 
incorporated into the analysis of conservation efforts in this final 
rule.
    We received information regarding the role of local management 
actions and conservation efforts with regard to reef resilience. 
Conservation projects and programs such as international agreements and 
memoranda of understanding, coral reef monitoring, voluntary protected 
areas, restoration activities, and outreach and education initiatives, 
among others, play an integral role in building and maintaining 
resilience within coral reef ecosystems as well as raising public 
awareness. More detailed information regarding local actions as they 
relate to reef resilience are described above in the Threats 
Evaluation, Inadequacy of Existing Regulatory Mechanisms section of 
this final rule.
    As described above, we received supplemental information about 
local conservation efforts since the publication of the proposed rule. 
However, we did not receive any supplemental information that changes 
our previous conclusions regarding global conservation efforts to slow 
climate change-related impacts. After considering this supplemental 
information in addition to that which was available for the proposed 
rule, our conclusions regarding conservation efforts remain unchanged. 
Overall, the numerous coral reef conservation projects are increasing 
and strengthening resiliency within coral reef ecosystems on a local 
level, and can provide a protective temporal buffer for corals in the 
face of climate change related threats. Coral reef restoration 
activities, particularly of the Caribbean acroporid species, are 
expected to assist in recovery efforts, but they cannot be considered a 
panacea. In the absence of effective global efforts to reduce impacts 
from climate change, there are no conservation efforts currently or 
planned in the future that are expected to improve the overall status 
of any of the listed species in this final rule, such that the 
additional protections provided by the ESA are not warranted.

Species-Specific Information and Determinations

Introduction

    This section summarizes the best available information for each of 
the 65

[[Page 53915]]

species of coral considered in this final rule. The best available 
information is comprised of the proposed rule and its supporting 
documents, and information that we either gathered ourselves or 
received as a result of public comments. To distinguish between the 
information on which the proposed rule was based from new or 
supplemental information, we will only cite the primary literature for 
new or supplemental information. For clarity, we will distinguish 
whether the information was identified via public comment or if we 
gathered it ourselves.
    Spatial, demographic, and other relevant biological information, 
threat susceptibilities, and information on regulatory mechanisms are 
all presented for each species. Because species-specific information is 
limited for many of the proposed species, genus-level information is 
highly relevant to our determinations. Therefore, we provide relevant 
information for each genus prior to providing the specific information 
for species within that genus. Specifically, genus-level information on 
threat susceptibilities is relevant to species when the available 
genus-level information can be appropriately applied to the species. 
Therefore, in each genus description, we provide a section that 
summarizes genus-level threat susceptibility information that was 
provided in the SRR and SIR, as well as in the public comments and 
supplemental information. Threat susceptibility conclusions are then 
provided considering the applicability of the genus-level information 
to an unstudied species within that genus. These conclusions will be 
applied, as appropriate, in the appropriate species descriptions.

Caribbean Species Determinations

Genus Agaricia

Introduction
    There are seven species in the genus Agaricia, all of which occur 
in the Caribbean (Veron, 2000). Colonies are composed of plates, which 
are flat, horizontal, or upright. The latter are usually contorted and 
fused. Some species such as A. humilis and Agaricia fragilis tend to be 
small and somewhat circular in shape while others like Agaricia 
lamarcki and Agaricia grahamae can form large, plating colonies.
Spatial Information
    The SRR and SIR provided the following genus-level information on 
Agaricia's distribution, habitat, and depth range: Agaricia can be 
found at depths of 50 to 100 m on mesophotic reefs.
    The public comments did not provide any new or supplemental 
information on Agaricia's distribution, habitat, and depth range. 
Supplemental information we found includes the following. Bongaerts et 
al. (2013) studied the depth distribution and genetic diversity of five 
agariciid species (A. humilis, A. agaricites, A. lamarcki, A. grahamae, 
and Helioseris cucullata [= Leptoseris cucullata]) and their symbiotic 
zooxanthellae in Cura[ccedil]ao. They found a distinct depth 
distribution among the species. Agaricia humilis and A. agaricites were 
more common at shallow depths, and A. lamarcki, A. grahamae, and H. 
cucullata were more common at deeper depths. They also found genetic 
segregation between coral host-symbiont communities at shallow and 
mesophotic depths.
Demographic Information
    The SRR and SIR provided the following genus level information on 
Agaricia's abundance and population trends: Coral specimens collected 
in 2010 from a mesophotic reef at Pulley Ridge, Florida suggest that 
corals, such as Agaricia spp., that appear live in video images may 
actually be covered with algae rather than live coral tissue.
    The public comments did not provide any new or supplemental 
information on Agaricia's abundance or population trends. Supplemental 
information we found on Agaricia's population trends includes the 
following: Stokes et al. (2010) reported a decrease in cover of 
Agaricia spp. in the Netherlands Antilles between 1982 and 2008 at all 
depths surveyed (10 to 30 m). An analysis of Caribbean monitoring data 
from 1970 to 2012 found that large, plating Agaricia spp. were one of 
the species groups that suffered the greatest proportional losses 
(Jackson et al., 2014).
Other Biological Information
    The SRR and SIR provided the following information on the life 
history of the genus Agaricia. In general, Agaricia spp. are gonochoric 
brooders. Several species such as Agaricia agaricites, A. tenuifolia, 
and A. humilis are known to use chemical cues from crustose coralline 
algae to mediate settlement.
    The public comments did not provide new or supplemental information 
on the life history of the genus Agaricia. Supplemental information we 
found on Agaricia's life history includes the following: Agaricia spp. 
can be one of the dominant taxonomic groups found in recruitment 
studies (Bak and Engel, 1979; Rogers et al., 1984; Shearer and 
Coffroth, 2006).
Susceptibility to Threats
    The SRR and SIR did not provide any genus level information on the 
susceptibility of Agaricia to ocean warming, and the public comments 
did not provide any new or supplemental information. Supplemental 
information we found on the susceptibility of the genus Agaricia to 
ocean warming includes the following: Agaricia is considered highly 
susceptible to bleaching. Agaricia spp. were the most susceptible to 
bleaching of the corals monitored during an unanticipated bleaching 
event at a remote, uninhabited island (Navassa), with higher bleaching 
prevalence at deeper sites (Miller et al., 2011a). During the 1998 
bleaching event in Belize, A. tenuifolia, a dominant coral, was nearly 
eradicated from the Channel Cay reef complex (Aronson et al., 2002). 
During the 2005 bleaching event, nearly all Agaricia spp. were bleached 
at long-term monitoring sites in Buck Island National Monument, and 
they remained bleached comparatively longer than other species 
monitored (Clark et al., 2009). Manzello et al. (2007) characterized 
Agaricia as having high susceptibility to bleaching in their study 
identifying bleaching indices and thresholds in the Florida Reef Tract, 
the Bahamas, and St. Croix, U.S. Virgin Islands. A long-term study in 
the Florida Keys found that bleaching prevalence was increased four to 
seven times by nutrient-enrichment in Agaricia spp., the only genus 
that showed such a response (Vega Thurber et al., 2014). This study 
indicated that the temperature threshold for bleaching may have been 
lowered by the nutrient enrichment. Notably, after removal of the 
nutrient enrichment, bleaching prevalence returned to background 
levels. Thus, we conclude that, absent species-specific information, 
species in the genus Agaricia should be considered highly susceptible 
to ocean warming-induced bleaching.
    The SRR and SIR did not provide any genus level information on the 
susceptibility of Agaricia to disease, and the public comments did not 
provide any new or supplemental information. Supplemental information 
we found on the susceptibility of the genus Agaricia to disease 
includes the following. A study of coral diseases across the wider-
Caribbean during the summer and fall of 2005 found the genus Agaricia, 
along with seven other major reef-building genera, to be particularly 
susceptible to coral diseases including white plague type II, Caribbean 
ciliate infection, and

[[Page 53916]]

to be infected with multiple diseases at the same time (Croquer and 
Weil, 2009). Agaricia agaricites decreased 87 percent in mean cover 
from the disease outbreak following the 2005 bleaching event in the 
U.S. Virgin Islands (Miller et al., 2009). Thus, we conclude that, 
absent species-specific information, species in the genus Agaricia 
should be considered highly susceptible to diseases.
    The SRR and SIR provided the following information on the 
susceptibility of Agaricia to acidification. No specific research has 
addressed the effects of acidification on the genus Agaricia. However, 
most corals studied have shown negative relationships between 
acidification and growth, and acidification is likely to contribute to 
reef destruction in the future. While ocean acidification has not been 
demonstrated to have caused appreciable declines in coral populations 
so far, it is considered a significant threat to corals by 2100.
    The public comments did not provide any new or supplemental 
information on the susceptibility of Agaricia to acidification. 
Supplemental information we found on the susceptibility of the genus 
Agaricia to acidification includes the following. Crook et al. (2012) 
surveyed coral populations near submarine springs close to the 
Mesoamerican Reef in Mexico where water aragonite saturation state was 
naturally low due to groundwater seepage. Agaricia spp. were found near 
the springs, but only in waters with an aragonite saturation state 
greater than 2.5, indicating these species may be less tolerant than 
other coral species that were able to grow in under-saturated waters. 
Thus, we conclude that, absent species-specific information, species in 
the genus Agaricia should be considered to have some susceptibility to 
acidification.
    The SRR and SIR provided genus level information on the 
susceptibility of Agaricia to sedimentation. The typically small 
calices of Agaricia spp. are not efficient at rejecting sediment, and 
species with horizontally-oriented plates or encrusting morphologies 
could be more sediment-susceptible than species with vertically-
oriented plates as evidenced by fine sediment suspended in hurricanes 
that caused higher mortality in platy corals than hemispherical or non-
flat ones. The public comments did not provide any new or supplemental 
information on the susceptibility of the genus Agaricia to 
sedimentation, and we did not find any new or supplemental information. 
Thus, we conclude that, absent species-specific information, species in 
the genus Agaricia should be considered to have some susceptibility to 
sedimentation.
    The SRR and SIR did not provide any genus level information on the 
susceptibility of Agaricia to nutrients, and the public comments did 
not provide any new or supplemental information. Supplemental 
information we found on the susceptibility of Agaricia spp. to 
nutrients includes the following. Treatment of A. tenuifolia with low 
(5 mg per l) and high (25 mg per l) doses of organic carbon resulted in 
73 to 77 percent mortality, respectively, compared to 10 percent 
mortality of controls (Kuntz et al. 2005). Treatment of A. tenuifolia 
with nitrate (7.5 [mu]M), ammonium (25 [mu]M), and phosphate (2.5 
[mu]M) caused about 50 percent mortality compared to 10 percent in 
controls (Kuntz et al. 2005). Thus, we conclude that, absent species-
specific information, species in the genus Agaricia should be 
considered to have high susceptibility to nutrient enrichment based on 
this study in combination with the Vega Thurber et al. (2014) study 
that found increased bleaching in the presence of chronic nutrient 
enrichment.
    The SRR and SIR did not provide any information on the 
susceptibility of Agaricia spp. to any other threats. The public 
comments did not provide any new or supplemental information, and we 
did not find any new or supplemental information on the susceptibility 
of Agaricia to any other threats.
Genus Conclusion
    The studies cited above indicate that Agaricia spp. are highly 
susceptible to warming. In at least one location, a bleaching event 
resulted in 100 percent mortality of one Agaricia species. The genus 
also appears to be highly susceptible to diseases that can result in 
high rates of mortality and to be highly susceptible to impacts of 
nutrients. However, as described below, there is a fair amount of 
species-specific information for individual Agaricia species; 
therefore, we generally do not rely on the genus-level information to 
inform species level determinations. When necessary the appropriate 
inference is described in the species-specific information.

Agaricia lamarcki

Introduction
    The SRR and SIR provided the following information on A. lamarcki's 
morphology and taxonomy. Agaricia lamarcki has flat, unifacial, or 
encrusting plates that are commonly arranged in whorls. It is 
identifiable by its morphology and the presence of white stars at the 
mouths. Agaricia lamarcki does not appear to have taxonomic problems.
    The public comments did not provide new or supplemental 
information, and we did not find any new or supplemental information on 
A. lamarcki's morphology or taxonomy.
Spatial Information
    The SRR and SIR provided the following information on A. lamarcki's 
distribution, habitat, and depth range. Agaricia lamarcki can be found 
in the western Atlantic off south Florida as far north as Palm Beach 
County, in the Gulf of Mexico including the Flower Garden Banks, and 
throughout the Caribbean including the Bahamas. Agaricia lamarcki is 
rare in shallow reef environments of 3 to 15 m, but is common at deeper 
depths of 20 to 100 m where it can be one of the dominant coral 
species. It is found in shaded or reduced light environments, on slopes 
and walls, and on mesophotic reefs in Cura[ccedil]ao, Florida, Jamaica, 
Puerto Rico, and the U.S. Virgin Islands.
    The public comments did not provide new or supplemental information 
on A. lamarcki's distribution, habitat, or depth range. Supplemental 
information we found on A. lamarcki's distribution includes the 
following. Veron (2014) confirms the presence of A. lamarcki in seven 
out of 11 possible ecoregions in the western Atlantic and greater 
Caribbean that contain corals, and he strongly predicts the presence of 
A. lamarcki in the ecoregion surrounding the Flower Garden Banks based 
on published record or confirmed occurrence in surrounding ecoregions. 
The three ecoregions in which it is not reported are off the coasts of 
Bermuda, Brazil, and the southeast U.S. north of south Florida. We did 
not find any new or supplemental information on A. lamarcki's habitat 
or depth range.
Demographic Information
    The SRR and SIR provided the following information on A. lamarcki's 
abundance and population trends. Agaricia lamarcki is reported as 
common. In the Netherlands Antilles, A. lamarcki increased in abundance 
or remained stable on reefs 30 to 40 m in depth from 1973 to 1992.
    The public comments provided supplemental information on A. 
lamarcki's abundance. Population estimates of A. lamarcki in the 
Florida Keys extrapolated from stratified random samples were 3.1 
 1.3 million (standard error (SE)) colonies in 2005 and 0.2 
 0.2 million colonies in 2012. No colonies were observed in 
2009, but

[[Page 53917]]

fewer deep sites (>20 m) were surveyed in 2009 and 2012 compared to 
2005. Most colonies observed were 20 to 30 cm in diameter, and partial 
mortality was highest (50 percent) in the largest size class (30 to 40 
cm). Agaricia lamarcki ranked 35th in abundance out of 47 species in 
2005 and 37th out of 40 species in 2012. In the Dry Tortugas, Florida, 
where more deep sites were surveyed, A. lamarcki ranked 12th out of 43 
species in 2006, with population estimates extrapolated to 14.3  2.6 million colonies. It ranked 22nd out of 40 species in 2008 
with populations estimates extrapolated to 2.1  0.5 million 
colonies. Most of the colonies in 2006 were 10 to 30 cm in diameter, 
but colonies greater than 90 cm were observed. Partial mortality was 
highest in the 30 to 40 cm size class (approximately 35 percent) in 
2006 and highest in the 20 to 30 cm size class (approximately 20 
percent) in 2008. In 2008, most of the colonies were 0 to 10 cm in 
size, and the largest colonies observed were in the 50 to 60 cm size 
class (Miller et al., 2013). Because population estimates were 
extrapolated from random samples, differences in population numbers 
between years are more likely a function of sampling effort rather than 
population trends over time. The public comments did not provide new or 
supplemental information on A. lamarcki's population trends.
    Supplemental information we found on A. lamarcki's abundance and 
population trends includes the following. Between 1977 and 1987, 
colonies of A. lamarcki in monitored plots in Jamaica decreased from 34 
to 31 colonies, indicating the net production by sexual and asexual 
means was not enough to compensate for mortality of the originally 
present colonies (Hughes, 1988). More than 40 percent of the colonies 
present in 1987 were derived from asexual fission of the original 
colonies present in 1977, and none of the six sexual recruits survived 
until the end of the study period (Hughes, 1988). In the U.S. Virgin 
Islands, A. lamarcki was the eleventh most common coral in terms of 
cover out of 55 species, and average cover across 18 monitoring sites 
was 1.2  0.3 (SE) percent in 2012 (Smith, 2013).
    All information on A. lamarcki's abundance and population trends 
can be summarized as follows. Based on population estimates, there are 
at least tens of millions of A. lamarcki colonies present in the 
Florida Keys and Dry Tortugas combined. Absolute abundance is higher 
than the estimate from these two locations given the presence of this 
species in many other locations throughout its range. Population trends 
indicate this species may be declining in some areas, but because some 
of the trend data is lumped by genus or genus plus morphology, there is 
uncertainty that the trends represent A. lamarcki specifically. Thus, 
we conclude that A. lamarcki has likely declined in some areas and the 
population numbers at least in the tens of millions of colonies.
Other Biological Information
    The SRR and SIR provided the following information on A. lamarcki's 
life history. No information on the reproductive strategy of A. 
lamarcki is available, but congeners are gonochoric brooders. Larval 
settlement occurs primarily at deeper depths (26 to 37 m), but the 
species has also been found at shallower depths. Recruitment rates of 
A. lamarcki are low (e.g., only one of 1,074 Agaricia recruits at the 
Flower Garden Banks may have been A. lamarcki), and net gains from 
sexual recruitment may be negligible at a decadal time scale. 
Population numbers may be maintained through asexual fission of larger 
colonies into smaller daughter colonies. Growth rates are slow; radial 
growth measurements from Jamaica ranged from zero to 1.4 cm per year 
and averaged approximately 0.5 cm per year. Growth rates are a bit 
slower, ranging from zero to 1.0 cm per year, at depths greater than 20 
m. Maximum colony size is approximately two meters. Agaricia lamarcki 
is a relatively long-lived species, and individual colonies may persist 
for greater than a century. Based on monitoring in Jamaica, the half-
life (mortality of half of monitored colonies) of A. lamarcki is 17 
years. Mortality rates are size-specific (ranging from 10 to 25 
percent), and partial mortality rates are high (ranging from 22 to 90 
percent). Overall, demographic characteristics are low recruitment, 
high colony survival, and high partial mortality.
    The public comments did not provide new or supplemental information 
on A. lamarcki life history. Supplemental information we found on A. 
lamarcki life history includes the following. Darling et al. (2012) 
performed a trait-based analysis to categorize coral species into four 
life history strategies: Generalist, weedy, competitive, and stress-
tolerant. The classifications were primarily separated by colony 
morphology, growth rate, and reproductive mode. Agaricia lamarcki was 
classified as a ``weedy'' species, thus likely more tolerant of 
environmental stress.
    The SRR, SIR, and the public comments did not provide new or 
supplemental biological information for A. lamarcki. Supplemental 
biological information we found about A. lamarcki includes the 
following. Out of five agariciid species sampled at a single reef in 
Cura[ccedil]ao, A. lamarcki was the only species that harbored multiple 
symbiont profiles across depth distribution; the other four species had 
only a single symbiont profile across depth. The symbiont community 
associated with A. lamarcki at 40 m depth was significantly different 
from those at both 10 m and 25 m (Bongaerts et al., 2013).
Susceptibility to Threats
    The threat susceptibility information from the SRR and SIR was 
interpreted in the proposed rule for A. lamarcki's vulnerabilities to 
threats as follows: Moderate vulnerability to ocean warming, disease, 
acidification, trophic effects of fishing, sedimentation, and 
nutrients; and low vulnerability to sea level rise and collection and 
trade. No conclusions on A. lamarcki's vulnerability to predation were 
made due to lack of available information on its susceptibility to this 
threat.
    The SRR and SIR provided the following information on the 
susceptibility of A. lamarcki to ocean warming. Agaricia lamarcki is 
susceptible to bleaching from both high and low temperature anomalies. 
In laboratory studies, A. lamarcki had almost complete disruption of 
photosynthesis at 32 [deg]C to 34 [deg]C. Bleaching can be extensive; 
however, it may not result in mortality in A. lamarcki.
    Van Woesik et al. (2012) developed a coral resiliency index to 
evaluate extinction risk due to bleaching, based on biological traits 
and processes. Evaluations were performed at the genus level. They 
rated the resiliency of Agaricia as -2 out of a range of -6 to 7 
observed in other coral genera. Less than or equal to -3 was considered 
highly vulnerable to extinction, and greater than or equal to 4 was 
considered highly tolerant. Thus, Agaricia was rated closer to the 
vulnerable end of the spectrum, though not highly vulnerable. This 
study was in the SIR, but the findings specific to Agaricia were not 
included. The public comments (comment 47) indicated the results of 
this study should be considered in the listing status of A. lamarcki.
    The public comments did not provide any new or supplemental 
information on the susceptibility of A. lamarcki to ocean warming. 
Supplemental information we found on the susceptibility of A. lamarcki 
to ocean warming includes the following. During the 2005 bleaching 
event, greater than

[[Page 53918]]

80 percent of A. lamarcki colonies bleached at 12 sites in Puerto Rico 
(Waddell and Clarke, 2008). In the U.S. Virgin Islands, an average of 
59 percent of A. lamarcki colonies (n = 11) bleached, and nine percent 
paled during the 2010 bleaching event (Smith et al., 2013b). Agaricia 
lamarcki had high resistance to both hot and cold water anomalies that 
impacted the Florida Keys in 2005 and 2010, respectively, as indicated 
by their low tissue mortality compared to other coral species monitored 
(Lirman et al., 2011).
    All sources of information are used to describe A. lamarcki's 
susceptibility to ocean warming as follows. Agaricia lamarcki has some 
susceptibility to ocean warming as evidenced by extensive bleaching 
during warm water temperature anomalies but observed low bleaching-
related mortality. The available information does not support a more 
precise description of susceptibility.
    The SRR and SIR did not provide any species-specific information on 
susceptibility of A. lamarcki to ocean acidification. The public 
comments did not provide new or supplemental information on the 
susceptibility of A. lamarcki to acidification, and we did not find any 
new or supplemental information.
    All sources of information are used to describe A. lamarcki's 
susceptibility to acidification as follows. There is uncertainty about 
how A. lamarcki will respond to ocean acidification, but there is 
genus-level evidence that Agaricia are not among the more tolerant 
species from areas of water with naturally lower aragonite saturation 
state. Thus, A. lamarcki likely has some susceptibility to ocean 
acidification, but the available information does not support a more 
precise description of susceptibility.
    The SRR and SIR provided the following information on A. lamarcki's 
susceptibility to disease. White plague infections in A. lamarcki have 
been observed in Florida, Colombia, and St. Lucia, though no incidence 
of disease was observed in the Florida Keys in 1996 to 1998. Ciliate 
infections have been documented in A. lamarcki, and tumors may affect 
this species. The ecological and population impacts of disease have not 
been established for A. lamarcki.
    The public comments did not provide any new or supplemental 
information on the susceptibility of A. lamarcki to disease, and we did 
not find any new or supplemental information on A. lamarcki's 
susceptibility to disease.
    All source of information are used to describe A. lamarcki's 
susceptibility to disease as follows. Agaricia lamarcki is susceptible 
to several diseases, including white plague, which has one of the 
fastest progression rates recorded in the Caribbean. However, there is 
no information on the population level effects of disease on A. 
lamarcki (e.g., rates of infection, percentage of population affected, 
and amounts of tissue loss). Genus-level information indicates high 
susceptibility to a disease outbreak following a bleaching event, 
indicating A. lamarcki is likely highly susceptible to disease.
    The SIR and SRR did not provide any species-specific information on 
the trophic effects of fishing on A. lamarcki. The public comments did 
not provide new or supplemental information, and we did not find new or 
supplemental information on the trophic effects of fishing on A. 
lamarcki. However, due to the level of reef fishing conducted in the 
Caribbean, coupled with Diadema die-off and lack of significant 
recovery, competition with algae can adversely affect coral 
recruitment. Thus, A. lamarcki likely has some susceptibility to the 
trophic effects of fishing because of low recruitment rates, though the 
available information does not support a more precise description of 
susceptibility.
    The SRR and SIR provided the following information on 
susceptibility of A. lamarcki to sedimentation. Agaricia lamarcki could 
be susceptible to sedimentation based on calix and colony morphology. 
This conclusion was based on genus-level information on susceptibility 
to sedimentation. The public comments did not provide new or 
supplemental information on the susceptibility of A. lamarcki to 
sedimentation, and we did not find new or supplemental information.
    All sources of information are used to describe A. lamarcki's 
susceptibility to sedimentation as follows. There is no species-
specific information on the susceptibility of A. lamarcki to 
sedimentation. However, based on genus-level information, colony 
morphology and skeletal structure of A. lamarcki indicate it is likely 
poor at removing sediment. Thus, A. lamarcki likely has some 
susceptibility to sedimentation, but the available information does not 
support a more precise description of susceptibility.
    The SRR and SIR did not provide any information on the 
susceptibility of A. lamarcki to nutrients, and the public comments did 
not provide any new or supplemental information. Supplemental 
information we gathered at the genus-level indicates that A. lamarcki 
is likely highly susceptible to nutrient enrichment.
    The SRR and SIR did not provide species-specific information on the 
effects of sea level rise on A. lamarcki. The SRR described sea level 
rise as an overall low to medium threat for all coral species. The 
public comments did not provide new or supplemental information on A. 
lamarcki's susceptibility to sea level rise, and we did not find any 
new or supplemental information. Thus, we conclude that A. lamarcki has 
some susceptibility to sea level rise, but the available information 
does not provide a more precise description of susceptibility.
    The SRR and SIR provided the following information on the 
susceptibility of A. lamarcki to collection and trade. Only light trade 
has been recorded with gross exports averaging fewer than 10 pieces of 
coral annually between 2000 and 2005. The public comments did not 
provide new or supplemental information on the susceptibility of A. 
lamarcki to collection and trade. Supplemental information we found 
confirms that collection and trade of A. lamarcki remained low between 
2000 and 2012 with gross exports averaging fewer than 10 pieces of 
coral annually (data available at http://trade.cites.org/). Thus, we 
conclude that A. lamarcki has low susceptibility to collection and 
trade.
    The SRR and SIR provided the following information on the 
susceptibility of A. larmarcki to predation. Predation effects on A. 
lamarcki are unknown. The public comments did not provide any new or 
supplemental information, and we did not find any new or supplemental 
information on the susceptibility of A. lamarcki to predation. We 
conclude that while A. lamarcki likely has some susceptibility to 
predation, available information is lacking, and we cannot say whether 
it is a threat.
Regulatory Mechanisms
    In the proposed rule, we relied on information from the Final 
Management Report for evaluating the existing regulatory mechanisms for 
controlling threats to all corals. However, we did not provide any 
species-specific information on the regulatory mechanisms or 
conservation efforts for A. lamarcki. Public comments were critical of 
that approach, and we therefore attempt to analyze regulatory 
mechanisms and conservation efforts on a species basis, where possible, 
in this final rule. Records confirm that Agaricia lamarcki occurs in 
eight Atlantic ecoregions that encompass 26 kingdom's and countries' 
EEZs. The 26 kingdoms and countries are Antigua &

[[Page 53919]]

Barbuda, Bahamas, Barbados, Belize, Colombia, Costa Rica, Cuba, 
Dominica, Dominican Republic, French Antilles, Grenada, Guatemala, 
Haiti, Kingdom of the Netherlands, Honduras, Jamaica, Mexico, 
Nicaragua, Panama, St. Kitts & Nevis, St. Lucia, St. Vincent & 
Grenadines, Trinidad and Tobago, United Kingdom (British Overseas 
Territories), United States (including U.S. Caribbean Territories), and 
Venezuela. The regulatory mechanisms relevant to A. lamarcki, described 
first as a percentage of the above kingdoms and countries that utilize 
them to any degree, and second as a percentage of those countries and 
kingdoms whose regulatory mechanisms may be limited in scope, are as 
follows: General coral protection (31 percent with 12 percent limited 
in scope), coral collection (50 percent with 27 percent limited in 
scope), pollution control (31 percent with 15 percent limited in 
scope), fishing regulations on reefs (73 percent with 50 percent 
limited in scope), managing areas for protection and conservation (88 
percent with 31 percent limited in scope). The most common regulatory 
mechanisms in place for A. lamarcki are reef fishing regulations and 
area management for protection and conservation. However, half of the 
reef fishing regulations are limited in scope and may not provide 
substantial protection for the species. General coral protection and 
collection laws, along with pollution control laws, are much less 
common regulatory mechanisms for the management of A. lamarcki.
Vulnerability to Extinction
    As explained above in the Risk Analyses section, a species' 
vulnerability to extinction results from the combination of its spatial 
and demographic traits, threat susceptibilities, and consideration of 
the baseline environment and future projections of threats. The SRR 
stated that the factors that increase the extinction risk for A. 
lamarcki include the widespread decline in environmental conditions in 
the Caribbean, potential losses to disease, severe effects of 
bleaching, and limited sediment tolerance. Factors that reduce 
extinction risk include occurrence primarily at great depth, where 
disturbance events are less frequent, and life history characteristics 
that have allowed the species to remain relatively persistent compared 
to other deep corals despite low rates of sexual recruitment.
    Subsequent to the proposed rule, we received and gathered 
supplemental species- or genus-specific information, described above, 
that expands our knowledge regarding the species' abundance, 
distribution, and threat susceptibilities. We developed our assessment 
of the species' vulnerability to extinction using all the available 
information. As explained in the Risk Analyses section, our assessment 
in this final rule emphasizes the ability of the species' spatial and 
demographic traits to moderate or exacerbate its vulnerability to 
extinction, as opposed to the approach we used in the proposed rule, 
which emphasized the species' susceptibility to threats.
    The following characteristics of A. lamarcki, in conjunction with 
the information described in the Corals and Coral Reefs section, Coral 
Habitat sub-section, and Threats Evaluation section above, affect its 
vulnerability to extinction currently and over the foreseeable future. 
Although it is geographically located in the heavily disturbed 
Caribbean, A. lamarcki's predominant occurrence at depths of 20 to 100 
m reduces its exposure to disturbance events that have resulted in the 
decreased resilience of reefs in the Caribbean and moderates 
vulnerability to extinction over the foreseeable future. Agaricia 
lamarcki's life history characteristics of large colony size and long 
life span have enabled it to remain relatively persistent despite slow 
growth and low recruitment rates, thus moderating vulnerability to 
extinction. Although we concluded that A. lamarcki is likely highly 
susceptible to disease, population level effects of disease have not 
been documented in A. lamarcki thus far, indicating the currently low 
vulnerability to extinction from this threat. Additionally, although A. 
lamarcki has been observed to have high levels of warming-induced 
bleaching, bleaching-related mortality appears to be low, indicating 
that vulnerability to extinction from ocean warming is currently low. 
Deeper areas of A. lamarcki's range will usually have lower 
temperatures than surface waters, and acidification is generally 
predicted to accelerate most in waters that are deeper and cooler than 
those in which the species occurs. Agaricia lamarcki's habitat includes 
shaded or reduced light environments, slopes, walls, and mesophotic 
reefs. This moderates vulnerability to extinction over the foreseeable 
future because the species is not limited to one habitat type but 
occurs in numerous types of reef environments that are predicted, on 
local and regional scales, to experience highly variable thermal 
regimes and ocean chemistry at any given point in time. Agaricia 
lamarcki's absolute abundance has been estimated as at least tens of 
millions of colonies in the Florida Keys and Dry Tortugas combined and 
is higher than the estimate from these two locations due to the 
occurrence of the species in many other areas throughout its range. Its 
abundance, life history characteristics, and depth distribution, 
combined with spatial variability in ocean warming and acidification 
across the species' range, moderate vulnerability to extinction because 
the increasingly severe conditions expected in the foreseeable future 
will be non-uniform, and there will likely be a large number of 
colonies that are either not exposed or do not negatively respond to a 
threat at any given point in time.
Listing Determination
    In the proposed rule, using the determination tool formula 
approach, A. lamarcki was proposed for listing as threatened because 
of: Moderate vulnerability to ocean warming (E), disease (C), and 
acidification (E); low relative recruitment rate (E); moderate overall 
distribution (based on narrow geographic distribution and wide depth 
distribution (E); restriction to the Caribbean (E); and inadequacy of 
regulatory mechanisms (D).
    In this final rule, we changed the listing determination for A. 
lamarcki from threatened to not warranted. We made this determination 
based on a more species-specific and holistic assessment of whether 
this species meets the definition of either a threatened or endangered 
coral, including more appropriate consideration of the buffering 
capacity of this species' spatial and demographic traits to lessen its 
vulnerability to threats. Thus, based on the best available information 
above on A. lamarcki' spatial structure, demography, threat 
susceptibilities, and management none of the five ESA listing factors, 
alone or in combination, are causing this species to be likely to 
become endangered throughout its range within the foreseeable future, 
and thus it is not warranted for listing at this time because:
    (1) Agaricia lamarcki's predominant occurrence at depths of 20 to 
100 m in heterogeneous habitats, including shaded or reduced light 
environments, on slopes and walls, and on mesophotic reefs, throughout 
the Caribbean basin reduces exposure to any given threat event or 
adverse condition that does not occur uniformly throughout the species' 
range. As explained above in the Threats Evaluation section, we have 
not identified any threat that is expected to occur uniformly 
throughout the species range within the foreseeable future; and

[[Page 53920]]

    (2) Agaricia lamarcki's absolute abundance is at least tens of 
millions of colonies based on estimates from two locations. Absolute 
abundance is higher than estimates from these locations since it occurs 
in many other locations throughout its range. This provides buffering 
capacity in the form of absolute numbers of colonies and variation in 
susceptibility between individual colonies. As discussed in the Corals 
and Coral Reefs section above, the more colonies a species has, the 
lower the proportion of colonies that are likely to be exposed to a 
particular threat at a particular time, and all individuals that are 
exposed will not have the same response.
    Notwithstanding the projections through 2100 that indicate 
increased severity over time of the three high importance threats, the 
combination of these biological and environmental characteristics 
indicates that the species possesses sufficient buffering capacity to 
avoid being in danger of extinction within the foreseeable future 
throughout its range. It is possible that this species' extinction risk 
may increase in the future if global threats continue and worsen in 
severity, and the species' exposure to the threats increases throughout 
its range. Should the species experience reduced abundance or range 
constriction of a certain magnitude, the ability of these 
characteristics to moderate exposure to threats will diminish. However, 
A. lamarcki is not likely to become of such low abundance or so 
spatially fragmented as to be in danger of extinction due to 
depensatory processes, the potential effects of environmental 
stochasticity, or the potential for mortality from catastrophic events 
within the foreseeable future throughout its range. Therefore, A. 
lamarcki is not warranted for listing at this time under any of the 
listing factors, and we withdraw our proposal to list the species as 
threatened.
Genus Mycetophyllia
    There are five species in the genus Mycetophyllia that all occur in 
the western Atlantic and Caribbean (Veron, 2000). Most species of 
Mycetophyllia can be difficult to distinguish in the field, and many 
studies report data to the genus level rather than species. Therefore, 
all information reported for the genus appears in this section, and 
information reported specifically for M. ferox is presented in the 
species section.
Demographic Information
    The SRR, SIR, and the public comments did not provide information 
on Mycetophyllia abundance or population trends. Supplemental 
information we found on Mycetophyllia's abundance and population trends 
includes the following. Percent cover of Mycetophyllia spp. between 
2001 and 2006 was less than approximately 0.02 percent on St. John (233 
sites surveyed) and St. Croix (768 sites surveyed), U.S. Virgin Islands 
and La Parguera, Puerto Rico (Waddell and Clarke, 2008). Similarly, 
cover of Mycetophyllia spp. on the mesophotic Hind Bank in the U.S. 
Virgin Islands was 0.02  0.01 percent in 2007 (Smith et 
al., 2010). Cover of Mycetophyllia spp. was 0.1 percent between 2002 
and 2004 on four islands in the Bahamas Archipelago (Roff et al., 
2011). Between 2005 and 2007, Mycetophyllia spp. comprised 0.1 percent 
or less of the coral cover and occurred in densities of 1.0 colony per 
10 m\2\ in parts of southeast Florida and the Florida Keys (Wagner et 
al., 2010). In Roatan, Honduras, Mycetophyllia sp. cover in permanent 
photo-stations increased between 1996 and 1998 from 0.57 percent to 
0.77 percent but subsequently decreased to 0.26 percent in 2003 and 
0.15 percent in 2005 (Riegl et al., 2009).
Susceptibility to Threats
    The SRR, SIR, and public comments did not provide information on 
Mycetophyllia's susceptibility to threats. Supplemental information we 
found on Mycetophyllia's susceptibility to ocean warming includes the 
following. During the 1995 bleaching event in Belize, 24 percent of 21 
colonies monitored Mycetophyllia bleached (McField, 1999). In Roatan, 
Honduras, 11 percent [sic]of 10 monitored Mycetophyllia sp. colonies 
bleached and 11 percent [sic] partially bleached during the 1998 
bleaching event; mortality of Mycetophyllia colonies was 11 percent 
(Riegl et al., 2009).
    Bleaching of Mycetophyllia was 62 percent across all 28 locations 
surveyed in Puerto Rico during the 2005 temperature anomaly (Waddell 
and Clarke, 2008). Additionally, a post-bleaching outbreak of white 
plague resulted in a massive collapse of Mycetophyllia colonies at most 
reefs on the east, south, and west coasts of Puerto Rico and 
reproductive failure during the 2006 mass spawning (Waddell and Clarke, 
2008). Off Mona and Desecheo Islands, Puerto Rico in 2005, paling 
occurred in 65 percent of Mycetophyllia colonies, and bleaching 
occurred in 10 percent (Bruckner and Hill, 2009).
    In surveys conducted between August and October 2005 to 2009 from 
the lower Florida Keys to Martin County, average mortality of 
Mycetophyllia spp. was 0.6  6.4 percent, which was the 
eighth highest out of 25 of the most abundant species (Lirman et al., 
2011). During the 2010 cold-water event, average mortality of 
Mycetophyllia spp. across 76 sites from the lower Florida Keys to 
Martin County was 15.0  28.3 percent, which was the 
eleventh highest of the 25 most abundant species (Lirman et al., 2011).
    During the 2005 bleaching event, Mycetophyllia spp. were among the 
most severely affected of 22 coral species reported to have bleached 
across 91 of 94 sites in northeast St. Croix, U.S. Virgin Islands 
(Wilkinson and Souter, 2008). In the U.S. Virgin Islands, the one 
colony of Mycetophyllia sp. observed at 18 sites, bleached during 2005. 
Six colonies were subsequently monitored after the 2010 mild bleaching 
event with average of eight percent bleaching (Smith et al., 2013b).
    Supplemental information we found on the susceptibility of 
Mycetophyllia to disease includes the following. White plague (Nugues, 
2002) and red band disease (Waddell, 2005) have been reported to infect 
Mycetophyllia species. In 2004, prevalence of disease in Mycetophyllia 
was approximately two to three percent in Mexico (Harvell et al., 
2007).

Mycetophyllia ferox

Introduction
    The SRR and SIR provided the following information on M. ferox's 
morphology and taxonomy. Mycetophyllia ferox forms a thin, encrusting 
plate that is weakly attached. Mycetophyllia ferox is taxonomically 
distinct. Maximum colony size is 50 cm.
    Public comments did not provide new or supplemental information on 
M. ferox's taxonomy or morphology. Supplemental information we found on 
M. ferox's taxonomy and morphology includes the following. Zlatarski 
and Estalella (1982) reported 14 out of 25 Mycetophyllia colonies 
collected from Cuba were intermediate between M. ferox, and M. 
lamarkiana, and parts of two colonies were comparable to M. ferox or M. 
lamarkiana, illustrating potential morphological plasticity between 
species.
Spatial Information
    The SRR and SIR provided the following information on M. ferox's 
distribution, habitat, and depth range. Mycetophyllia ferox occurs in 
the western Atlantic and throughout the wider Caribbean. It has not 
been reported in the Flower Garden Banks (Gulf of Mexico) or in 
Bermuda. It has been reported in reef environments in

[[Page 53921]]

water depths of 5 to 90 m, including shallow and mesophotic habitats.
    The public comments did not provide new or supplemental information 
on M. ferox's distribution, habitat, or depth range. Supplemental 
information we found on M. ferox's distribution includes the following. 
Veron (2014) confirms the occurrence of M. ferox in seven out of a 
possible 11 ecoregions in the Caribbean and western Atlantic that 
contain corals. The four ecoregions where it is not reported are the 
Flower Garden Banks, off the coasts of Bermuda, Brazil, and the 
southeast U.S. north of south Florida. We did not find any supplemental 
information on M. ferox's habitat or depth range.
Demographic Information
    The SRR and SIR provided the following information on M. ferox's 
abundance and population trends. Mycetophyllia ferox is usually 
uncommon or rare, constituting less than 0.1 percent of all coral 
species at generally less than one percent of the benthic cover. 
Density of M. ferox in southeast Florida and the Florida Keys was 
approximately 0.8 colonies per 10 m\2\ between 2005 and 2007. There is 
indication that the species was much more abundant in the upper Florida 
Keys in the 1970s. In a survey of 97 stations in the Florida Keys, M. 
ferox declined in occurrence from 20 stations in 1996 to four stations 
in 2009. At 21 stations in the Dry Tortugas, M. ferox declined in 
occurrence from eight stations in 2004 to three stations in 2009.
    The public comments provided the following supplemental information 
on M. ferox's abundance. In stratified random surveys in the Florida 
Keys, M. ferox ranked 39th most abundant out of 47 in 2005, 43rd out of 
43 in 2009, and 40th out of 40 in 2012. Extrapolated population 
estimates were 1.0  0.7 (SE) million in 2005, 9,500  9,500 (SE) colonies in 2009, and 7,000  7,000 (SE) 
in 2012 . These abundance estimates are based on random surveys, and 
differences between years are more likely a result of sampling effort 
rather than population trends. The most abundant size class was 10 to 
20 cm diameter that equaled the combined abundance of the other size 
classes. The largest size class was 30 to 40 cm. Average partial 
mortality per size class ranged from nearly 0 to 50 percent and was 
greatest in the 20 to 30 cm size class (Miller et al., 2013).
    In the Dry Tortugas, Florida, M. ferox ranked 35th most abundant 
out of 43 species in 2006 and 30th out of 40 in 2008. Population 
estimates were 0.5  0.4 (SE) million in 2006 and 0.5  0.2 million (SE) in 2008. The number of colonies in 2006 was 
similar between the 0 to 10 cm and 10 to 20 cm size classes, and the 
largest colonies were in the 20 to 30 cm size class. Greatest partial 
mortality was around 10 percent. Two years later, in 2008, the highest 
proportion of colonies was in the 20 to 30 cm size class, and the 
largest colonies were in the 40 to 50 cm size class. The greatest 
partial mortality was about 60 percent in the 30 to 40 cm size class, 
however the number of colonies at that size were few (Miller et al., 
2013).
    Supplemental information we found on M. ferox's abundance and 
population trends confirms M. ferox's low percent cover, encounter 
rate, and density. In a survey of Utila, Honduras between 1999 and 
2000, M. ferox was observed at eight percent of 784 surveyed sites and 
was the 36th most commonly observed out of 46 coral species; other 
Mycetophyllia species were seen more commonly (Afzal et al., 2001). In 
surveys of remote southwest reefs of Cuba, M. ferox was observed at one 
of 38 reef-front sites, with average abundance was 0.004  
0.027 (standard deviation (SD)) colonies per 10 m transect; this was 
comparatively lower than the other three Mycetophyllia species observed 
(Alcolado et al., 2010). Between 1998 and 2004, cover of M. ferox 
ranged between 0.3 and 0.4 percent in three of six sites monitored in 
Colombia (Rodriguez-Ramirez et al., 2010). In Barbados, M. ferox was 
observed on one of seven reefs surveyed, and the average cover was 0.04 
percent (Tomascik and Sander, 1987).
    Benthic cover of M. ferox in the Red Hind Marine Conservation 
District off St. Thomas, U.S. Virgin Islands, which includes mesophotic 
coral reefs, was 0.003  0.004 percent in 2007, accounting 
for 0.02 percent of coral cover, and ranking 20th highest in cover out 
of 21 coral species (Nemeth et al., 2008; Smith et al., 2010). In the 
U.S. Virgin Islands between 2001 and 2012, cover of M. ferox appeared 
in 12 of 33 survey sites and accounted for 0.01 percent of the benthos, 
and 0.07 percent of the coral cover, ranking as 13th most common 
(Smith, 2013).
    In 1981, M. ferox was observed on one of four reefs surveyed in the 
upper Florida Keys at 0.1 percent cover (Burns, 1985). In surveys of 
the Florida Keys between 1996 and 2003, cover of M. ferox was 0.022, 
0.005, and less than 0.001 percent on patch reefs, deep offshore reefs, 
and shallow offshore reefs, respectively (Somerfield et al., 2008). At 
permanent monitoring stations in the Florida Keys, the number of 
stations where M. ferox was present declined between 1996 and 2003 
(Waddell, 2005). Between 2005 and 2010, M. ferox was one of 42 species 
surveyed and was found the least abundant being observed at densities 
of 0.02 and 0.01 colonies per 10 m\2\ on mid-channel reefs and fore-
reefs, respectively, on the Florida reef tract (Burman et al., 2012).
    All information on M. ferox's abundance and population trends can 
be summarized as follows. Mycetophyllia ferox has been reported to 
occur on 3 to 50 percent of reefs surveyed and is one of the least 
common coral species observed. On reefs where M. ferox is found, it 
generally occurs at abundances of less than one colony per 10 m\2\ and 
percent cover of less than 0.1 percent. Based on population estimates, 
there are at least hundreds of thousands of M. ferox colonies present 
in the Florida Keys and Dry Tortugas combined. Absolute abundance is 
higher than the estimate from these two locations given the presence of 
this species in many other locations throughout its range. Low 
encounter rate and percent cover coupled with the tendency to include 
Mycetophyllia spp. at the genus level make it difficult to discern 
population trends of M. ferox from monitoring data. However, reported 
losses of M. ferox from monitoring stations in the Florida Keys and Dry 
Tortugas (63 to 80 percent loss) indicate population decline in these 
locations. Based on declines in Florida, we conclude M. ferox has 
likely declined throughout its range.
Other Biological Information
    The SRR and SIR provided the following information on M. ferox's 
life history. Mycetophyllia ferox is a hermaphroditic brooding species. 
Colony size at first reproduction is greater than 100 cm\2\. 
Recruitment of M. ferox appears to be very low, even in studies from 
the 1970s.
    The public comments did not provide new or supplemental information 
on M. ferox's life history. Supplemental information we found on M. 
ferox's life history includes the following. Mycetophyllia ferox has a 
lower fecundity compared to M. aliciae, M. lamarckiana and M. danaana 
(Morales Tirado, 2006). Over a 10 year period, no colonies of M. ferox 
were observed to recruit to an anchor-damaged site in the U.S. Virgin 
Islands although adults were observed on the adjacent reef (Rogers and 
Garrison, 2001). Darling et al. (2012) performed a biological trait-
based analysis to categorize coral species into four life history 
strategies: Generalist, weedy, competitive, and stress-tolerant. 
Mycetophyllia ferox was classified as a

[[Page 53922]]

``weedy'' species, thus likely more tolerant of environmental stress.
Susceptibility to Threats
    The threat susceptibility information from the SRR and SIR was 
interpreted in the proposed rule for M. ferox's vulnerabilities to 
threats as follows: High vulnerability to disease and nutrient 
enrichment; moderate vulnerability to ocean warming, acidification, 
trophic effects of fishing, and sedimentation; and low vulnerability to 
sea level rise, predation, and collection and trade.
    The SRR and SIR provided the following information on M. ferox's 
susceptibility to ocean warming. No bleached M. ferox colonies were 
observed in Florida or Barbados in a wide-scale survey during the 2005 
mass-bleaching event, although the number of colonies was small.
    The public comments did not provide new or supplemental information 
on the susceptibility of M. ferox to ocean warming. Supplemental 
information we found on the susceptibility of M. ferox to ocean warming 
includes the following. In surveys of the lower Florida Keys and Dry 
Tortugas during the 1998 bleaching event, approximately 20 percent of 
M. ferox colonies bleached; out of the 14 species reported to have 
experienced bleaching of at least 50 percent of the colony, M. ferox 
was one of the least affected (Waddell, 2005). Approximately 50 percent 
of M. ferox colonies bleached at 12 locations in Puerto Rico during the 
2005 bleaching event (Waddell and Clarke, 2008). During the 2005 
Caribbean bleaching event, neither of the two colonies of M. ferox 
monitored at six sites in Barbados bleached; an average of 71 percent 
of all coral colonies bleached at those six sites during the event 
(Oxenford et al., 2008).
    All sources of information are used to describe M. ferox's 
susceptibility to ocean warming as follows. The bleaching reports 
available specifically for M. ferox and at the genus level indicate 
similar trends of relatively low bleaching observed in 1995, 1998, and 
2010 (less than 25 percent) and higher levels (50 to 65) or no 
bleaching in the more severe 2005 bleaching event. Reproductive failure 
and a disease outbreak were reported for the genus after the 2005 
bleaching event. Although bleaching of most coral species is spatially 
and temporally variable, understanding the susceptibility of M. ferox 
is somewhat confounded by the species' low sample size in any given 
survey due to its low encounter rate. We conclude that M. ferox has 
some susceptibility to ocean warming. However, the available 
information does not support a more precise description of 
susceptibility to this threat.
    The SRR and SIR provided the following information on the 
susceptibility of M. ferox to acidification. No specific research has 
addressed the effects of acidification on the genus Mycetophyllia. 
However, most corals studied have shown negative relationships between 
acidification and growth, and acidification is likely to contribute to 
reef destruction in the future. While ocean acidification has not been 
demonstrated to have caused appreciable declines in coral populations 
to date, it is considered to become a significant threat to corals by 
2100.
    The public comments did not provide new or supplemental information 
on the susceptibility of M. ferox to acidification, and we did not find 
any new or supplemental information.
    All sources of information are used to describe M. ferox's 
susceptibility to acidification as follows. There is uncertainty about 
how M. ferox will respond to ocean acidification. Based on the negative 
effects of acidification on growth of most corals, M. ferox likely has 
some susceptibility to acidification. The available information does 
not support a more precise description of susceptibility.
    The SRR and SIR provided the following information on M. ferox's 
susceptibility to disease. Mycetophyllia ferox is susceptible to white 
plague. Diseased M. ferox colonies were reported in the upper Florida 
Keys in the mid-1970s; between 24 and 73 percent of M. ferox colonies 
were infected per site. At one reef site, 20 to 30 percent of the M. 
ferox colonies died from disease during a one-year period.
    The public comments did not provide new or supplemental information 
on the susceptibility of M. ferox to disease. Supplemental information 
we found on the susceptibility of M. ferox to disease includes the 
following. Porter et al. (2001) report the loss of M. ferox from many 
of the permanent monitoring stations (160 stations at 40 sites) in the 
Florida Keys between 1996 and 1998 due to coral disease.
    All sources of information are used to describe M. ferox's 
susceptibility to disease as follows. From reports in the Florida Keys, 
M. ferox appears to be highly susceptible to disease, specifically 
white plague, and reports of high losses and correlation with higher 
temperatures date back to the mid-1970s (Dustan, 1977). Although heavy 
impacts of disease on M. ferox have not been reported in other 
locations, an outbreak of white plague was credited with causing heavy 
mortality at the genus level in Puerto Rico after the 2005 bleaching 
event. We conclude that the susceptibility of M. ferox to disease is 
high.
    The SIR and SRR did not provide any species-specific information on 
the trophic effects of fishing on M. ferox. The public comments did not 
provide new or supplemental information, and we did not find new or 
supplemental information on the trophic effects of fishing on M. ferox. 
However, due to the level of reef fishing conducted in the Caribbean, 
coupled with Diadema die-off and lack of significant recovery, 
competition with algae can adversely affect coral recruitment. Thus, M. 
ferox likely has some susceptibility to the trophic effects of fishing 
given its low recruitment rates. The available information does not 
support a more precise description of susceptibility.
    The SRR and SIR provided the following information on the 
susceptibility of M. ferox to nutrient enrichment. Mycetophyllia ferox 
appeared to be absent at fringing reef sites in Barbados impacted by 
sewage pollution.
    The public comments did not provide any new or supplemental 
information on the susceptibility of M. ferox to nutrient enrichment, 
and we did not find any new or supplemental information.
    All sources of information are used to describe M. ferox's 
susceptibility to nutrient enrichment as follows. Mycetophyllia ferox 
may be susceptible to nutrient enrichment as evidenced by its absence 
from eutrophic sites in one location. However, there is uncertainty 
about whether the absence is a result of eutrophic conditions or a 
result of uncommon or rare occurrence. Therefore, we conclude that M. 
ferox likely has some susceptibility to nutrient enrichment. However, 
the available information does not support a more precise description 
of susceptibility.
    The SRR and SIR did not provide any species or genus information on 
the susceptibility of M. ferox to sedimentation but provided the 
following. Land-based sources of pollution (including sediment) often 
act in concert rather than individually and are influenced by other 
biological (e.g., herbivory) and hydrological factors. Collectively, 
land-based sources of pollution are unlikely to produce extinction at a 
global scale; however, they may pose significant threats at local 
scales and reduce the resilience of corals to bleaching.
    The public comments did not provide new or supplemental information 
on the

[[Page 53923]]

susceptibility of M. ferox to sedimentation, and we did not find any 
new or supplemental information. We conclude that M. ferox has some 
level of susceptibility to sedimentation, but the available information 
does not support a more precise description of susceptibility.
    The SRR and SIR provided the following information on the 
susceptibility of M. ferox to predation. Mycetophyllia ferox has not 
been susceptible to predation. Public comments did not provide new or 
supplemental information on M. ferox's susceptibility to predation, and 
we did not find any new or supplemental information. We conclude that 
M. ferox has low susceptibility to predation.
    The SRR and SIR did not provide species-specific information on the 
effects of sea level rise on M. ferox. The SRR described sea level rise 
as an overall low to medium threat for all coral species. The public 
comments did not provide new or supplemental information on M. ferox's 
susceptibility to sea level rise, and we did not find any new or 
supplemental information. Thus, we conclude that M. ferox has some 
susceptibility to sea level rise, but the available information does 
not provide a more precise description of susceptibility.
    The SRR and SIR provided the following information on M. ferox's 
susceptibility to collection and trade. Mycetophyllia ferox is not 
reported to be an important species for trade. Exports of M. ferox were 
ten pieces in 2000, two in 2003, and five in 2007.
    The public comments did not provide new or supplemental information 
on the susceptibility of M. ferox to collection and trade. Supplemental 
information we found confirmed low collection and trade of M. ferox 
with gross exports between 2000 and 2012 averaging fewer than two 
corals per year (data available at http://trade.cites.org/). Thus, we 
conclude that M. ferox has low susceptibility to collection and trade.
Regulatory Mechanisms
    In the proposed rule, we relied on information from the Final 
Management Report for evaluating the existing regulatory mechanisms for 
controlling threats to all corals. However, we did not provide any 
species-specific information on the regulatory mechanisms or 
conservation efforts for M. ferox. Public comments were critical of 
that approach, and we therefore attempt to analyze regulatory 
mechanisms and conservation efforts on a species basis, where possible, 
in this final rule. Records confirm that M. ferox occurs in seven 
Atlantic ecoregions that encompass 26 kingdom's or countries' EEZs. The 
26 kingdoms and countries are Antigua & Barbuda, Bahamas, Barbados, 
Belize, Colombia, Costa Rica, Cuba, Dominica, Dominican Republic, 
French Antilles, Grenada, Guatemala, Haiti, Honduras, Jamaica, Kingdom 
of the Netherlands, Mexico, Nicaragua, Panama, St. Kitts & Nevis, St. 
Lucia, St. Vincent & Grenadines, Trinidad and Tobago, United Kingdom 
(British Overseas Territories), United States (including U.S. Caribbean 
Territories), and Venezuela. The regulatory mechanisms relevant to M. 
ferox, described first as a percentage of the above kingdoms and 
countries that utilize them to any degree, and, second as the 
percentages of those kingdoms and countries whose regulatory mechanisms 
may be limited in scope, are as follows general coral protection (31 
percent with 12 percent limited in scope), coral collection (50 percent 
with 27 percent limited in scope), pollution control (31 percent with 
15 percent limited in scope), fishing regulations on reefs (73 percent 
with 50 percent limited in scope), managing areas for protection and 
conservation (88 percent with 31 percent limited in scope). The most 
common regulatory mechanisms in place for M. ferox are reef fishing 
regulations and area management for protection and conservation. 
However, half of the reef-fish fishing regulations are limited in scope 
and may not provide substantial protection for the coral species. 
General coral protection and collection laws, along with pollution 
control laws, are much less common regulatory mechanisms for the 
management of M. ferox.
Vulnerability to Extinction
    As explained above in the Risk Analyses section, a species' 
vulnerability to extinction results from the combination of its spatial 
and demographic traits, threat susceptibilities, and consideration of 
the baseline environment and future projections of threats. The SRR 
stated that the factors that increase the extinction risk for M. ferox 
include disease, rare abundance, and observed declines in abundance.
    Subsequent to the proposed rule, we received and gathered 
supplemental species- or genus-specific information, described above, 
that expands our knowledge regarding the species' abundance, 
distribution, and threat susceptibilities. We developed our assessment 
of the species' vulnerability to extinction using all the available 
information. As explained in the Risk Analyses section, our assessment 
in this final rule emphasizes the ability of the species' spatial and 
demographic traits to moderate or exacerbate its vulnerability to 
extinction, as opposed to the approach we used in the proposed rule, 
which emphasized the species' susceptibility to threats.
    The following characteristics of M. ferox, in conjunction with the 
information described in the Corals and Coral Reefs section, Coral 
Habitat sub-section, and Threats Evaluation section above, affect its 
vulnerability to extinction currently and over the foreseeable future. 
Mycetophyllia ferox has declined due to disease in at least a portion 
of its range and has low recruitment, which limits its capacity for 
recovery from mortality events and exacerbates vulnerability to 
extinction. Despite the large number of islands and environments that 
are included in the species' range, geographic distribution in the 
highly disturbed Caribbean exacerbates vulnerability to extinction over 
the foreseeable future because M. ferox is limited to an area with 
high, localized human impacts and predicted increasing threats. Its 
depth range of five to 90 meters moderates vulnerability to extinction 
over the foreseeable future because deeper areas of its range will 
usually have lower temperatures than surface waters, and acidification 
is generally predicted to accelerate most in waters that are deeper and 
cooler than those in which the species occurs. Its habitat includes 
shallow and mesophotic reefs which moderates vulnerability to 
extinction over the foreseeable future because the species occurs in 
numerous types of reef environments that are predicted, on local and 
regional scales, to experience highly variable thermal regimes and 
ocean chemistry at any given point in time. Mycetophyllia ferox is 
usually uncommon to rare throughout its range. Its absolute abundance 
has been estimated as at least hundreds of thousands of colonies in the 
Florida Keys and Dry Tortugas combined and is higher than the estimate 
from these two locations due to the occurrence of the species in many 
other areas throughout its range. Its abundance, combined with spatial 
variability in ocean warming and acidification across the species' 
range, moderate vulnerability to extinction because the threats are 
non-uniform, and there will likely be a large number of colonies that 
are either not exposed or do not negatively respond to a threat at any 
given point in time.
Listing Determination
    In the proposed rule, using the determination tool formula 
approach, M. ferox was proposed for listing as endangered because of: 
High

[[Page 53924]]

vulnerability to disease (C); moderate vulnerability to ocean warming 
(E) and acidification (E); high vulnerability to nutrient over-
enrichment (A and E); rare general range-wide abundance (E); decreasing 
trend in abundance (E); low relative recruitment rate (E); moderate 
overall distribution (based on narrow geographic distribution and wide 
depth distribution (E); restriction to the Caribbean (E); and 
inadequacy of regulatory mechanisms (D).
    In this final rule, we changed the listing determination for M. 
ferox from endangered to threatened. We made this determination based 
on a more species-specific and holistic approach, including 
consideration of the buffering capacity of this species' spatial and 
demographic traits, and the best available information above on M. 
ferox's spatial structure, demography, threat susceptibilities, and 
management. This combination of factors indicates that M. ferox is 
likely to become endangered throughout its range within the foreseeable 
future, and thus warrants listing as threatened at this time, because:
    (1) Mycetophyllia ferox is highly susceptible to disease (C) and 
susceptible to ocean warming (ESA Factor E), acidification (E), trophic 
effects of fishing (A), nutrients (A, E), and sedimentation (A, E). 
These threats are expected to continue and increase into the future. In 
addition, the species is at heightened extinction risk due to 
inadequate existing regulatory mechanisms to address global threats 
(D);
    (2) Mycetophyllia ferox has experienced significant declines in 
Florida and has likely experienced decline in other locations in its 
range;
    (3) Mycetophyllia ferox has a usually uncommon to rare occurrence 
throughout its range, which heightens the potential effect of localized 
mortality events and leaves the species vulnerable to becoming of such 
low abundance within the foreseeable future that it may be at risk from 
depensatory processes, environmental stochasticity, or catastrophic 
events, as explained in more detail in the Corals and Coral Reefs and 
Risk Analyses sections;
    (4) Mycetophyllia ferox is geographically located in the highly 
disturbed Caribbean where localized human impacts are high and threats 
are predicted to increase as described in the Threats Evaluation 
section. A range constrained to this particular geographic area that is 
likely to experience severe and increasing threats indicates that a 
high proportion of the population of this species is likely to be 
exposed to those threats over the foreseeable future; and
    (5) Mycetophyllia ferox's low recruitment limits the capacity for 
recovery from threat-induced mortality events throughout the range over 
the foreseeable future.
    The combination of these characteristics and future projections of 
threats indicates that the species is likely to be in danger of 
extinction within the foreseeable future throughout its range and 
warrants listing as threatened at this time due to factors A, C, D, and 
E.
    The available information above on M. ferox's spatial structure, 
demography, threat vulnerabilities, and management also indicate that 
the species is not currently in danger of extinction and thus does not 
warrant listing as Endangered because:
    (1) While Mycetophyllia ferox's distribution within the Caribbean 
increases its risk of exposure to threats as described above, its depth 
distribution is five to 90 m and its habitat includes various shallow 
and mesophotic reef environments. This moderates vulnerability to 
extinction currently because the species is not limited to one habitat 
type but occurs in numerous types of reef environments that will 
experience highly variable thermal regimes and ocean chemistry on local 
and regional scales at any given point in time, as described in more 
detail in the Coral Habitat and Threats Evaluation sections. There is 
no evidence to suggest that the species is so spatially fragmented that 
depensatory processes, environmental stochasticity, or the potential 
for catastrophic events currently pose a high risk to the survival of 
the species; and
    (2) Mycetophyllia ferox's absolute abundance is at least hundreds 
of thousands of colonies based on estimates from two locations. 
Absolute abundance is higher than estimates from these locations since 
M. ferox occurs in many other locations throughout its range. This 
absolute abundance allows for variation in the responses of individuals 
to threats to play a role in moderating vulnerability to extinction for 
the species to some degree, as described in more detail in the Corals 
and Coral Reefs section. Its absolute abundance indicates it is 
currently able to avoid high mortality from environmental 
stochasticity, and mortality of a high proportion of its population 
from catastrophic events.
    The combination of these characteristics indicates that the species 
does not exhibit the characteristics of one that is currently in danger 
of extinction, as described previously in the Risk Analyses section, 
and thus does not warrant listing as endangered at this time.
    Range-wide, multitudes of conservation efforts are already broadly 
employed that are likely benefiting M. ferox. However, considering the 
global scale of the most important threats to the species, and the 
ineffectiveness of conservation efforts at addressing the root cause of 
global threats (i.e., GHG emissions), we do not believe that any 
current conservation efforts or conservation efforts planned in the 
future will result in affecting the species' status to the point at 
which listing is not warranted.

Genus Dendrogyra

    The SRR and SIR provided the following information on morphology 
and taxonomy of Dendrogyra. Dendrogyra cylindrus is the only species in 
the genus Dendrogyra. It is easily identifiable, and there is no 
taxonomic confusion. The public comments did not provide new or 
supplemental information on the morphology or taxonomy of D. cylindrus, 
and we did not find any new or supplemental information.

Dendrogyra cylindrus

Introduction
    The SRR and SIR provided the following information on the 
morphology of D. cylindrus. Dendrogyra cylindrus forms cylindrical 
columns on top of encrusting bases. Colonies are generally grey-brown 
in color and may reach three meters in height. Tentacles remain 
extended during the day, giving columns a furry appearance.
Spatial Information
    The SRR and SIR provided the following information on D. 
cylindrus's distribution, habitat, and depth range. Dendrogyra 
cylindrus is present in the western Atlantic and throughout the greater 
Caribbean. The SRR reports a single known colony in Bermuda that is in 
poor condition. Dendrogyra cylindrus inhabits most reef environments in 
water depths ranging from one to 25 m.
    The public comments did not provide new or supplemental information 
on D. cylindrus's distribution, habitat, or depth range. Supplemental 
information we found on D. cylindrus's distribution, habitat, and depth 
range include the following. Dendrogyra cylindrus is absent from the 
southwest Gulf of Mexico (Tunnell, 1988). There is fossil evidence of 
the presence of D. cylindrus off Panama less than 1000 years ago, but 
it has been reported as absent today (Florida Fish and Wildlife 
Conservation Commission, 2013). Veron (2014)

[[Page 53925]]

confirms the presence of D. cylindrus in seven out of a potential 11 
ecoregions in the western Atlantic and wider-Caribbean that are known 
to contain corals. The four ecoregions in which it is not reported are 
the Flower Garden Banks and off the coasts of Bermuda, Brazil, and the 
southeast U.S. north of south Florida. Although D. cylindrus's depth 
range is 1 to 25 m, it is most common between five and 15 m depth 
(Acosta and Acevedo, 2006; Cairns, 1982; Goreau and Wells, 1967).
    All information on D. cylindrus's distribution can be summarized as 
follows. Dendrogyra cylindrus is distributed throughout most of the 
greater Caribbean in most reef environments between 1 to 25 m depth. It 
currently appears to be absent from Panama where it historically 
occurred within the last 1000 years.
Demographic Information
    The SRR and SIR provided the following information on D. 
cylindrus's abundance and population trends. Dendrogyra cylindrus is 
uncommon but conspicuous with scattered, isolated colonies. It is 
rarely found in aggregations. Dendrogyra cylindrus has been reported to 
be common on Pleistocene reefs around Grand Cayman, but rare on modern 
reefs. In monitoring studies, cover is generally less than one percent. 
Between 2005 and 2007, mean density of D. cylindrus was approximately 
0.5 colonies per 10 m\2\ in the Florida Keys. In a study of D. 
cylindrus demographics at Providencia Island, Colombia, a total of 283 
D. cylindrus colonies were detected in a survey of 1.66 km\2\ for and 
overall density of 172.0  177.0 (SE) colonies per km\2\.
    The public comments provided supplemental information on D. 
cylindrus's abundance but not on population trends. In stratified 
random samples of the Florida Keys, D. cylindrus ranked least common 
out of 47 coral species in 2005 and 41 out of 43 species in 2009. Based 
on random surveys stratified by habitat type, extrapolated abundance 
for the Florida Keys was 23,000  23,000 (SE) colonies in 
2005 and 25,000  25,000 (SE) colonies in 2009. Because 
these population estimates were based on random sampling, differences 
between years is more likely a function of sampling effort rather than 
an indication of population trends. All D. cylindrus colonies reported 
in 2005 were in the 70 to 80 cm diameter size class with less than two 
percent partial mortality. Four years later in 2009, all reported 
colonies were greater than 90 cm. No D. cylindrus colonies were 
encountered in 600 surveys from Key Biscayne to Key West, Florida in 
2012, with the authors noting sampling design was not optimized for 
this species. This species was not reported in the Dry Tortugas in 2006 
and 2008, and rarely encountered during pilot studies conducted over 
several years (1999 to 2002) ranking 49th out of 49 coral species 
(Miller et al., 2013).
    Supplemental information we found on D. cylindrus's abundance and 
population trends confirms the uncommon occurrence, rare encounter 
rate, low percent cover, and low density. During surveys of Utila, 
Honduras between 1999 and 2000, D. cylindrus was sighted in 19.6 
percent of 784 surveys and ranked 26th most common in abundance out of 
48 coral species (Afzal et al., 2001). In surveys of the upper Florida 
Keys in 2011, D. cylindrus was the second rarest out of 37 coral 
species and encountered at one percent of sites (Miller et al., 2011b).
    In stratified random surveys from Palm Beach County to the Dry 
Tortugas, Florida between 2005 and 2010, D. cylindrus was seen only on 
the ridge complex and mid-channel reefs at densities of 1.09 and 0.1 
colonies per 10 m\2\, respectively (Burman et al., 2012). Average 
number of D. cylindrus colonies in remote reefs off southwest Cuba was 
0.013  0.045 colonies per 10 m transect, and the species 
ranked sixth rarest out of 38 coral species (Alcolado et al., 2010).
    Out of 283 D. cylindrus colonies at Providencia Island, Colombia, 
70 were fragments resulting from asexual fragmentation, and no sexual 
recruits were observed. Size class distribution was skewed to smaller 
size classes less than 60 cm in height, and average colony height was 
73.8  46.0 cm (Acosta and Acevedo, 2006).
    Dendrogyra cylindrus's average percent cover was 0.002 on patch 
reefs and 0.303 in shallow offshore reefs in annual surveys of 37 sites 
in the Florida Keys between 1996 and 2003 (Somerfield et al., 2008). At 
permanent monitoring stations in the U.S. Virgin Islands, D. cylindrus 
has been observed in low abundance at 10 of 33 sites and, where 
present, ranged in cover from less than 0.05 percent to 0.22 percent 
(Smith, 2013). In Dominica, D. cylindrus comprised less than 0.9 
percent cover and was present at 13.3 percent of 31 surveyed sites 
(Steiner, 2003). At seven fringing reefs off Barbados, D. cylindrus was 
observed on one reef, and cover was 2.7  1.4 percent 
(Tomascik and Sander, 1987). In monitored photo-stations in Roatan, 
Honduras, cover of D. cylindrus increased slightly from 1.35 percent in 
1996 to 1.67 percent in 1999 and then declined to 0.44 percent in 2003 
and 0.43 percent in 2005 (Riegl et al., 2009). In the U.S. Virgin 
Islands, seven percent of 26 monitored colonies experienced total 
colony mortality between 2005 and 2007, though the very low cover of D. 
cylindrus (0.04 percent) remained relatively stable during this time 
period (Smith et al., 2013b).
    All sources of information on D. cylindrus's abundance and 
population trends can be summarized as follows. Based on population 
estimates, there are at least tens of thousands of D. cylindrus 
colonies present in the Florida Keys. Absolute abundance is higher than 
the estimate from this location given the presence of this species in 
many other locations throughout its range. Although there is evidence 
of potentially higher population levels in some areas of the Caribbean 
during the Pleistocence, D. cylindrus is currently uncommon to rare. 
Few studies report D. cylindrus population trends, and the low 
abundance and infrequent encounter rate in monitoring programs result 
in small samples sizes. The low coral cover of this species renders 
monitoring data difficult to extrapolate to realize trends. Therefore, 
we conclude that D. cylindrus is naturally uncommon to rare and that 
trends are unknown.
Other Biological Information
    The SRR and SIR provided the following information on D. 
cylindrus's life history. Dendrogyra cylindrus is a gonochoric 
(separate sexes) broadcast spawning species with relatively low annual 
egg production for its size. The combination of gonochoric spawning 
with persistently low population densities is expected to yield low 
rates of successful fertilization and low larval supply. Sexual 
recruitment of this species is low, and reported juvenile colonies in 
the Caribbean are lacking. Dendrogyra cylindrus can propagate by 
fragmentation following storms or other physical disturbance. Average 
growth rates of 1.8 to 2.0 cm per year in linear extension have been 
reported in the Florida Keys compared to 0.8 cm per year in Colombia 
and Cura[ccedil]ao. Partial mortality rates are size-specific with 
larger colonies having greater rates. Frequency of partial mortality 
can be high (65 percent of 185 colonies surveyed in Colombia), while 
the amount of partial mortality per colony is generally low (average of 
3 percent of tissue area affected per colony).
    The public comments did not provide new or supplemental information 
on D. cylindrus's life history. Supplemental information we found on D. 
cylindrus's life history includes the following. Spawning observations 
have been made

[[Page 53926]]

several nights after the full moon of August in the Florida Keys (Neely 
et al., 2013; Waddell and Clarke, 2008).
    Darling et al. (2012) performed a biological trait-based analysis 
to categorize coral species into four life history strategies: 
Generalist, weedy, competitive, and stress-tolerant. The 
classifications were primarily separated by colony morphology, growth 
rate, and reproductive mode. Dendrogyra cylindrus was classified as a 
``competitive'' species, thus likely more vulnerable to environmental 
stress.
    The SRR and SIR provided the following other biological information 
for D. cylindrus. Dendrogyra cylindrus appears to be sensitive to cold 
temperatures. Feeding rates (removal of suspended particles in 
seawater) are low relative to most other Caribbean corals, indicating 
it is primarily a tentacle feeder rather than a suspension feeder. 
However, D. cylindrus has a relatively high photosynthetic rate, and 
stable isotope values suggest it receives substantial amounts of 
photosynthetic products from its zooxanthellae.
    The public comments did not provide new or supplemental biological 
information for D. cylindrus. Supplemental information we found 
confirms that D. cylindrus is sensitive to cold temperatures and is 
summarized as follows. In laboratory studies of cold shock, D. 
cylindrus had the highest zooxanthellae expulsion rate of three species 
tested at 12 degrees C (Muscatine et al., 1991). During the 2010 cold 
water event in the Florida Keys, D. cylindrus was one of the most 
affected coral species with 100 percent mortality on surveyed inshore 
reefs (Kemp et al., 2011).
Susceptibility to Threats
    The threat susceptibility information from the SRR and SIR was 
interpreted in the proposed rule for D. cylindrus's vulnerabilities to 
threats as follows: High vulnerability to disease; moderate 
vulnerability to ocean warming, acidification, trophic effects of 
fishing, sedimentation, and nutrient enrichment; and low vulnerability 
to sea level rise, predation, and collection and trade.
    The SRR and SIR provided the following information on the 
susceptibility of D. cylindrus to ocean warming. There are conflicting 
characterizations of bleaching susceptibility of D. cylindrus in the 
literature. The species was bleaching-resistant during the 1983 mass-
bleaching event in Florida. Characterizations of the 2005 mass-
bleaching event in southern Florida and in the U.S. Virgin Islands 
noted that no bleached D. cylindrus colonies were observed, but during 
the same event in Barbados 100 percent of 15 D. cylindrus colonies 
bleached.
    Van Woesik et al. (2012) developed a coral resiliency index based 
on biological traits and processes to evaluate extinction risk due to 
bleaching. Evaluations were performed at the genus level. They rated 
the resiliency of D. cylindrus as 3 out of a range of -6 to 7 observed 
in other coral genera. Less than or equal to -3 was considered highly 
vulnerable to extinction, and greater than or equal to 4 was considered 
highly tolerant. Thus, D. cylindrus was rated as moderately tolerant. 
While this study was included in the SIR, species-specific findings for 
Dendrogyra were not included. The public comments (Comment 47) 
indicated the results of this study should be considered in the listing 
status of D. cylindrus.
    The public comments did not provide new or supplemental information 
on the susceptibility of D. cylindrus to ocean warming. Supplemental 
information we found confirms the variable susceptibility of D. 
cylindrus to ocean warming and bleaching. Dendrogyra cylindrus was 
among 42 species reported not to have bleached at various locations in 
the western Atlantic (British Virgin Islands, Jamaica, and Mona Island) 
during the 1987 bleaching event, while the authors noted these species 
were reported bleached at other locations or other areas by others 
(Williams and Bunkley-Williams, 1990). None of the 18 D. cylindrus 
colonies monitored in Roatan, Honduras experienced bleaching or 
mortality in the 1998 event where bleaching ranged from zero to 89 
percent in the 22 species monitored (Riegl et al., 2009). Across 12 
locations in Puerto Rico, 100 percent of D. cylindrus colonies bleached 
during the 2005 temperature anomaly (Waddell and Clarke, 2008). 
However, Bruckner and Hill (2009) report less severe D. cylindrus 
bleaching during the 2005 event in Puerto Rico; approximately 25 
percent paled and 10 percent bleached on reefs off Mona and Desecheo 
Islands, which was relatively low compared to some other species such 
as Orbicella faveolata, which had approximately 60 percent bleached 
colonies. At Dairy Bull Reef in Jamaica, 50 percent of D. cylindrus 
colonies bleached during the 2005 bleaching event, but no mortality was 
reported for this species (Quinn and Kojis, 2008). An average of 33 
percent of the monitored D. cylindrus colonies in the U.S. Virgin 
Islands bleached in 2005, and 67 percent paled. None of the monitored 
colonies bleached or paled during the less severe 2010 bleaching event 
(Smith et al., 2013b).
    All sources of information are used to describe D. cylindrus's 
susceptibility to ocean warming as follows. There are conflicting 
characterizations of the susceptibility of D. cylindrus to bleaching. 
Some locations experienced high bleaching of up to 100 percent of D. 
cylindrus colonies during the 2005 Caribbean bleaching event while 
others had a smaller proportion of colonies bleach (10 to 50 percent). 
Reports of low mortality after less severe bleaching indicate potential 
resilience, though mortality information is absent from locations that 
reported high bleaching frequency. Although bleaching of most coral 
species is spatially and temporally variable, understanding the 
susceptibility of D. cylindrus is further confounded by the species' 
rarity and, hence, low sample size in any given survey. We conclude 
that although D. cylindrus appears to have resistance to bleaching from 
warmer temperatures in some portions of its range under some 
circumstances, it is likely to have some susceptibility to ocean 
warming, given the high rates of bleaching observed at times. However, 
the available information does not support a more detailed description 
of susceptibility.
    The SRR and SIR provided the following information on the 
susceptibility of D. cylindrus to acidification. No specific research 
has addressed the effects of acidification on the genus Dendrogyra. 
However, most corals studied have shown negative relationships between 
acidification and growth, and acidification is likely to contribute to 
reef destruction in the future. While ocean acidification has not been 
demonstrated to have caused appreciable declines in coral populations 
so far, it is considered a significant threat to corals by 2100.
    The public comments did not provide new or supplemental information 
on the susceptibility of D. cylindrus to acidification, and we did not 
find any new or supplemental information.
    All sources of information are used to describe D. cylindrus's 
susceptibility to acidification as follows. Dendrogyra cylindrus likely 
has some susceptibility to acidification, but the available information 
does not support a more precise description of susceptibility to this 
threat.
    The SRR and SIR provided the following information on the 
susceptibility of D. cylindrus to disease. Dendrogyra cylindrus is 
susceptible to black band disease and white plague, though impacts from 
white plague are likely more extensive because of rapid progression 
rates. The large colony size suggests that individual colonies are less 
likely to suffer complete mortality

[[Page 53927]]

from a given disease exposure, but low colony density suggests that 
even small degrees of mortality increase extinction risk.
    The public comments did not provide new or supplemental information 
on the susceptibility of D. cylindrus to disease. Supplemental 
information we found on the susceptibility of D. cylindrus to disease 
includes the following. In a January 2002 survey at Providencia Island, 
Colombia, 4.2 percent of D. cylindrus colonies (n=185) exhibited white 
plague type II (Acosta and Acevedo, 2006). The prevalence of diseased 
D. cylindrus colonies was approximately three percent in Mexico from 
2002 to 2004 (Ward et al., 2006). Though white diseases were reported 
to cause colony mortality in some coral species in the U.S. Virgin 
Islands after the 2005 Caribbean bleaching event, none of the monitored 
D. cylindrus colonies exhibited signs of white disease (Smith et al., 
2013b).
    All sources of information are used to describe D. cylindrus's 
susceptibility to disease as follows. Disease appears to be present in 
about three to four percent of the population in some locations. 
Because no studies have tracked disease progression in D. cylindrus, 
the effects of disease are uncertain at both the colony and population 
level. However, the reported low partial mortality and large colony 
size suggest that individual colonies are less likely to suffer 
complete colony mortality from a given disease exposure. Therefore, we 
conclude that D. cylindrus has some susceptibility to disease, but the 
available information does not support a more precise description of 
susceptibility to this threat.
    The SIR and SRR did not provide any species-specific information on 
the trophic effects of fishing on D. cylindrus. The public comments did 
not provide new or supplemental information, and we did not find new or 
supplemental information on the trophic effects of fishing on D. 
cylindrus. However, due to the level of reef fishing conducted in the 
Caribbean, coupled with Diadema die-off and lack of significant 
recovery, competition with algae can adversely affect coral 
recruitment. This effect coupled with the species' low recruitment rate 
indicates it likely has some susceptibility to the trophic effects of 
fishing. The available information does not support a more precise 
description of its susceptibility.
    The SRR and SIR provided the following information on the 
susceptibility of D. cylindrus to sedimentation. The rate of sand 
removal from D. cylindrus tissues in laboratory conditions was 
intermediate among 19 Caribbean coral species tested.
    The public comments did not provide new or supplemental information 
on the susceptibility of D. cylindrus to sedimentation. Supplemental 
information we found includes the following. Dendrogyra cylindrus, 
along with Acropora spp. and Meandrina meandrites, was found in fossil 
assemblages only on the reef tract and not on the lagoonal patch reefs 
around Grand Cayman, suggesting that this species may be ineffective at 
sediment rejection like the other two species or may be intolerant of 
turbidity (Hunter and Jones, 1996).
    All sources of information are used to describe D. cylindrus's 
susceptibility to sedimentation as follows. Dendrogyra cylindrus 
appears to be moderately capable of removing sediment from its tissue. 
However, D. cylindrus may be more sensitive to turbidity due to its 
high reliance on nutrition from photosynthesis and as evidenced by the 
geologic record. Therefore, we conclude that D. cylindrus has some 
susceptibility to sedimentation, but the available information does not 
support a more precise description of susceptibility to this threat.
    The SRR and SIR provided the following information on the 
susceptibility of D. cylindrus to nutrient enrichment. Along a 
eutrophication gradient in Barbados, D. cylindrus was found at a single 
site, one of those farthest removed from pollution. The public comments 
did not provide new or supplemental on the susceptibility of D. 
cylindrus to nutrient enrichment, and we did not find any new or 
supplemental information.
    All sources of information are used to describe D. cylindrus's 
susceptibility to nutrient enrichment as follows. Dendrogyra cylindrus 
may be susceptible to nutrient enrichment as evidenced by its absence 
from eutrophic sites in one location. However, there is uncertainty 
about whether its absence is a result of eutrophic conditions or a 
result of its naturally uncommon or rare occurrence. Therefore, we 
conclude that D. cylindrus likely has some susceptibility to nutrient 
enrichment. However, the available information does not support a more 
precise description of its susceptibility to this threat.
    The SRR and SIR provided the following information on the 
susceptibility of D. cylindrus to predation. The corallivorous fireworm 
Hermodice carunculata has been observed feeding on diseased colonies of 
D. cylindrus, but generally, predation is not observed to cause 
noticeable mortality on D. cylindrus, despite its rarity.
    The public comments did not provide new or supplemental information 
on D. cylindrus's susceptibility to predation. Supplemental information 
we found includes the following. The sea urchin, Diadema antillarum, 
has been reported to cause partial mortality at the base of D. 
cylindrus colonies (Acosta and Acevedo, 2006).
    All sources of information are used to describe D. cylindrus's 
susceptibility to predation as follows. The low amounts of observed 
mortality indicate D. cylindrus has low susceptibility to predation.
    The SRR and SIR did not provide species-specific information on the 
effects of sea level rise on D. cylindrus. The SRR described sea level 
rise as an overall low to medium threat for all coral species. The 
public comments did not provide new or supplemental information on D. 
cylindrus's susceptibility to sea level rise, and we did not find any 
new or supplemental information. Thus, we conclude that D. cylindrus 
has some susceptibility to sea level rise, but the available 
information does not support a more precise description of 
susceptibility to this threat.
    The SRR and SIR provided information on D. cylindrus's 
susceptibility to collection and trade. Overall trade reports indicate 
very low rates of international trade of D. cylindrus. It is possible 
that historical curio collecting of D. cylindrus may have significantly 
reduced populations off Florida.
    The public comments did not provide new or supplemental information 
of the susceptibility of D. cylindrus to collection and trade. 
Supplemental information we found confirms what was provided by the SRR 
and SIR. Prior to its ban in the 1980s, collection of D. cylindrus for 
curios was once widespread off the coast of Florida (Florida Fish and 
Wildlife Conservation Commission, 2013). From 2000 to 2012, 
international trade of this species was low with gross exports ranging 
from zero to nine corals per year (average less than two per year; data 
available at http://trade.cites.org).
    All sources of information are used to describe D. cylindrus's 
susceptibility to collection and trade as follows. In the past, 
collection and trade may have had a large effect on the population in 
some locations like Florida. However, collection and trade likely does 
not have a large impact on the population currently. Therefore, we 
conclude that the susceptibility of D. cylindrus to collection and 
trade is currently low.

[[Page 53928]]

Regulatory Mechanisms
    In the proposed rule, we relied on information from the Final 
Management Report for evaluating the existing regulatory mechanisms for 
controlling threats to all corals. However, we did not provide any 
species-specific information on the regulatory mechanisms or 
conservation efforts for D. cylindrus. Public comments were critical of 
that approach, and we therefore attempt to analyze regulatory 
mechanisms and conservation efforts on a species basis, where possible, 
in this final rule. Records confirm that D. cylindrus occurs in seven 
Atlantic ecoregions that encompass 26 kingdom's and countries' EEZs. 
The 26 kingdoms and countries are Antigua & Barbuda, Bahamas, Barbados, 
Belize, Colombia, Costa Rica, Cuba, Dominica, Dominican Republic, 
French Antilles, Grenada, Guatemala, Haiti, Kingdom of the Netherlands, 
Honduras, Jamaica, Mexico, Nicaragua, Panama, St. Kitts & Nevis, St. 
Lucia, St. Vincent & Grenadines, Trinidad and Tobago, United Kingdom 
(British Caribbean Territories), United States (including U.S. 
Caribbean Territories), and Venezuela. The regulatory mechanisms 
relevant to D. cylindrus, described first as a percentage of the above 
kingdoms and countries that utilize them to any degree, and, second as 
the percentages of those kingdoms and countries whose regulatory 
mechanisms may be limited in scope, are as follows: General coral 
protection (31 percent with 12 percent limited in scope), coral 
collection (50 percent with 27 percent limited in scope), pollution 
control (31 percent with 15 percent limited in scope), fishing 
regulations on reefs (73 percent with 50 percent limited in scope), 
managing areas for protection and conservation (88 percent with 31 
percent limited in scope). The most common regulatory mechanisms in 
place for D. cylindrus are reef fishing regulations and area management 
for protection and conservation. However, half of the reef fishing 
regulations are limited in scope and may not provide substantial 
protection for the species. General coral protection and collection 
laws, along with pollution control laws, are much less common 
regulatory mechanisms for the management of D. cylindrus.
    Dendrogyra cylindrus is listed as threatened on the State of 
Florida endangered and threatened species list. The state has an action 
plan for conservation of the species with several objectives including 
stabilizing or increasing the existing population, the current area of 
occupancy, and the number of sexually mature individuals and evaluating 
the reproductive potential of the population over the next decade 
(Florida Fish and Wildlife Conservation Commission, 2013). However, the 
management plan recognizes that there are threats to D. cylindrus that 
need to be addressed outside the scope of the plan in order to improve 
the status of this species.
Vulnerability to Extinction
    As explained above in the Risk Analyses section, a species' 
vulnerability to extinction results from the combination of its 
demographic and spatial characteristics, threat susceptibilities, and 
consideration of the baseline environment and future projections of 
threats. The SRR stated that the factors that increase the extinction 
risk for D. cylindrus include the overall low population density and 
low population size, gonochoric spawning mode and lack of observed 
sexual recruitment, and susceptibility to observed disease mortality. 
The SRR acknowledged that, given the apparent naturally rare status of 
this species, some undescribed adaptations to low population density 
may exist in this species, particularly with regard to overcoming 
fertilization limitation between spawned gametes from gonochoric parent 
colonies that are at great distance from one another. Nonetheless, the 
pervasiveness of threats characterizing the Caribbean region was deemed 
to represent substantial extinction risk given this species' low 
population size.
    Subsequent to the proposed rule, we received and gathered 
supplemental species- or genus-specific information, described above, 
that expands our knowledge regarding the species' abundance, 
distribution, and threat susceptibilities. We developed our assessment 
of the species' vulnerability to extinction using all the available 
information. As explained in the Risk Analyses section, our assessment 
in this final rule emphasizes the ability of the species' spatial and 
demographic traits to moderate or exacerbate its vulnerability to 
extinction, as opposed to the approach we used in the proposed rule, 
which emphasized the species' susceptibility to threats.
    The following characteristics of D. cylindrus, in conjunction with 
the information described in the Corals and Coral Reefs section, Coral 
Habitat sub-section, and Threats Evaluation section above, affect its 
vulnerability to extinction currently and over the foreseeable future. 
Dendrogyra cylindrus is susceptible to a number of threats, but there 
is little evidence of population declines thus far. Despite the large 
number of islands and environments that are included in the species' 
range, geographic distribution in the highly disturbed Caribbean 
exacerbates vulnerability to extinction over the foreseeable future 
because D. cylindrus is limited to an area with high, localized human 
impacts and predicted increasing threats. Dendrogyra cylindrus inhabits 
most reef environments in water depths ranging from 1 to 25 m which 
moderates vulnerability to extinction over the foreseeable future 
because the species occurs in numerous types of reef environments that 
are predicted, on local and regional scales, to experience highly 
variable thermal regimes and ocean chemistry at any given point in 
time. It is naturally rare. Estimates of absolute abundance are at 
least tens of thousands of colonies in the Florida Keys, and absolute 
abundance is higher than estimates from this location due to the 
occurrence of the species in many other areas throughout its range. It 
is a gonochoric broadcast spawner with observed low sexual recruitment. 
Its low abundance, combined with its geographic location, exacerbates 
vulnerability to extinction because increasingly severe conditions 
within the species' range are likely to affect a high proportion of its 
population at any given point in time, and low sexual recruitment is 
likely to inhibit recovery potential from mortality events, further 
exacerbating its vulnerability to extinction.
Listing Determination
    In the proposed rule, using the determination tool formula 
approach, D. cylindrus was proposed for listing as endangered because 
of: High vulnerability to disease (C); moderate vulnerability to ocean 
warming (E) and acidification (E); rare general range-wide abundance 
(E); low relative recruitment rate (E); narrow overall distribution 
(based on narrow geographic distribution and moderate depth 
distribution (E); restriction to the Caribbean (E); and inadequacy of 
regulatory mechanisms (D).
    In this final rule, we changed the listing determination for D. 
cylindrus from endangered to threatened. We made this determination 
based on a more species-specific and holistic approach, including 
consideration of the buffering capacity of this species' spatial and 
demographic traits, and the best available information above on D. 
cylindrus's spatial structure, demography, threat susceptibilities, and 
management. This combination of factors indicates that D. cylindrus is 
likely to become endangered throughout

[[Page 53929]]

its range within the foreseeable future, and thus warrants listing as 
threatened at this time, because:
    (1) Dendrogyra cylindrus is susceptible to ocean warming (ESA 
Factor E), disease (C), acidification (E), nutrient enrichment (A and 
E), sedimentation (A and E), and trophic effects of fishing (A). These 
threats are expected to continue and increase into the future. In 
addition, the species is at heightened extinction risk due to 
inadequate existing regulatory mechanisms to address global threats 
(D).
    (2) Dendrogyra cylindrus is geographically located in the highly 
disturbed Caribbean where localized human impacts are high and threats 
are predicted to increase as described in the Threats Evaluation 
section. A range constrained to this particular geographic area that is 
likely to experience severe and increasing threats indicates that a 
high proportion of the population of this species is likely to be 
exposed to those threats over the foreseeable future;
    (3) Dendrogyra cylindrus has an uncommon to rare occurrence 
throughout its range, which heightens the potential effect of localized 
mortality events and leaves the species vulnerable to becoming of such 
low abundance within the foreseeable future that it may be at risk from 
depensatory processes, environmental stochasticity, or catastrophic 
events, as explained in more detail in the Corals and Coral Reefs and 
Risk Analyses sections; and
    (4) Dendrogyra cylindrus's low sexual recruitment limits its 
capacity for recovery from threat-induced mortality events throughout 
its range over the foreseeable future.
    The combination of these characteristics and future projections of 
threats indicates the species is likely to be in danger of extinction 
within the foreseeable future throughout its range and warrants listing 
as threatened at this time due to factors A, C, D, and E.
    The available information above on D. cylindrus spatial structure, 
demography, threat susceptibilities, and management also indicate that 
the species is not currently in danger of extinction and thus does not 
warrant listing as Endangered because:
    (1) There is little evidence of D. cylindrus population declines 
(i.e., the species continues to be naturally rare);
    (2) Dendrogyra cylindrus shows evidence of resistance to bleaching 
from warmer temperatures in some portions of its range under some 
circumstances (e.g., Roatan, Honduras); and
    (3) While D. cylindrus's distribution within the Caribbean 
increases its risk of exposure to threats as described above, its 
habitat includes most reef environments in water depths ranging from 
one to 25 m. This moderates vulnerability to extinction currently 
because the species is not limited to one habitat type but occurs in 
numerous types of reef environments that will experience highly 
variable thermal regimes and ocean chemistry on local and regional 
scales at any given point in time, as described in more detail in the 
Coral Habitat and Threats Evaluation sections.
    The combination of these characteristics indicates that the species 
does not exhibit the characteristics of one that is currently in danger 
of extinction, as described previously in the Risk Analyses section, 
and thus does not warrant listing as endangered at this time.
    Last, D. cylindrus is listed as threatened on the State of Florida 
endangered and threatened species list, and an action plan for 
conservation has recently been developed. Implementation of the action 
plan will no doubt have benefits to the species, but it is too soon to 
evaluate its effectiveness for conserving the species. Further, 
considering the global scale of the most important threats to the 
species, and the ineffectiveness of conservation efforts at addressing 
the root cause of global threats (i.e., GHG emissions), we do not 
believe that any current conservation efforts or conservation efforts 
planned in the future will result in affecting the species' status to 
the point at which listing is not warranted.

Genus Dichocoenia

    The SRR and SIR provided the following information on Dichocoenia's 
morphology and taxonomy. There are potentially two species in the genus 
Dichocoenia: Dichocoenia stokesi and Dichocoenia stellaris. Dichocoenia 
stellaris has been described as differing from D. stokesi by its 
pancake-like colony morphology and dominance of smaller, circular 
calices. Some coral taxonomists consider there to be only one species, 
D. stokesi, as specimens have all variations of skeletal shape and 
valley length. The public comments did not provide any new or 
supplemental information on Dichocoenia's taxonomy or morphology, and 
we did not find any new or supplemental information.
    Most studies over the last several decades describe D. stokesi and 
do not separately report data for colonies with D. stellaris 
morphology. Because D. stokesi was petitioned for listing and D. 
stellaris was not, we considered all information on D. stokesi and did 
not consider information on D. stellaris, despite some uncertainty of 
whether or not these are the same species. If D. stokesi is accepted to 
include all sizes of calices, it is easy to identify; if not then 
species delineations are somewhat arbitrary. We did not find any 
supplemental information on Dichocoenia's taxonomy.

Dichocoenia stokesi

Introduction
    Dichocoenia stokesi forms mounding-spherical colonies that are 
usually orange-brown but sometimes green.
Spatial Information
    The SRR and SIR provided the following information on D. stokesi's 
distribution, habitat, and depth range. Dichocoenia stokesi is located 
in the western Atlantic, Gulf of Mexico (including the Florida Middle 
Grounds and Flower Garden Banks), and throughout the Caribbean. It is 
also reported in Bermuda, though it is rare. Dichocoenia stokesi occurs 
in most reef environments within its range, including mesophotic reefs, 
back- and fore-reef environments, rocky reefs, lagoons, spur-and-groove 
formations, channels, and occasionally at the base of reefs. It has 
been reported in water depths ranging from two to 72 m.
    The public comments did not provide any new or supplemental 
information on D. stokesi's distribution, habitat, or depth range. 
Supplemental information we found includes the following. Veron (2014) 
confirmed the occurrence of D. stokesi in nine out of 11 ecoregions in 
the western Atlantic and wider-Caribbean known to contain corals. The 
two ecoregions in which it is not reported are off the coasts of 
Brazil, and the southeast U.S. north of south Florida. Kahng et al. 
(2010) report that D. stokesi is relatively abundant and dominates the 
coral community on mesophotic reefs greater than 40 m depth in the 
northern Gulf of Mexico but not in Belize, Puerto Rico, U.S. Virgin 
Islands, Jamaica, Curacao, Florida, Bermuda, Bahamas, or Barbados.
    All information on D. stokesi's distribution can be summarized as 
follows. Dichocoenia stokesi is distributed throughout most of the 
greater Caribbean in most reef environments within its range, including 
mesophotic reefs.
Demographic Information
    The SRR and SIR provided the following information on D. stokesi 
abundance. Dichocoenia stokesi is characterized as usually uncommon. In 
surveys of southeast Florida and the

[[Page 53930]]

Florida Keys between 2005 and 2007, D. stokesi comprised between 1.8 
and 7.0 percent of all coral colonies observed and was present at a 
density of approximately 1.7 colonies per 10 m\2\, which was the ninth 
most abundant out of an observed 43 coral species.
    The public comments provided the following supplemental information 
on D. stokesi's abundance. In stratified random surveys conducted by 
Miller et al. (2013) in the Florida Keys, D. stokesi ranked as the 8th 
most abundant species or higher in 2005, 2009, and 2012. Extrapolated 
abundance was 97.8  13.1 (SE) million colonies in 2005, 
53.8  9.7 (SE) million colonies in 2009, and 81.6  10.0 (SE) million colonies in 2012. Because population estimates 
were based on random sampling, differences between years are more 
likely a function of sampling effort rather than an indication of 
population trends. Most colonies were 30 cm or less in size, and size 
class distributions remained similar among the three sample periods 
(2005, 2009, and 2012). Larger colonies typically exhibited more 
partial mortality, which ranged between 20 and 80 percent for colonies 
larger than 10 cm.
    In the Dry Tortugas, D. stokesi was ranked 12th and 14th most 
common in 2006 and 2008, respectively. Extrapolated colony abundance 
was 12.1  4.1 (SE) million colonies in 2006 and 7.1  1.1 (SE) million colonies in 2008. All D. stokesi colonies 
observed were 40 cm or less in 2006, and 20 cm or less in 2008. Partial 
mortality was higher in larger colonies and ranged from approximately 
20 to 65 percent in colonies larger than 10 cm (Miller et al., 2013).
    Supplemental information we found on D. stokesi's abundance 
includes the following. In surveys of Utila, Honduras between 1999 and 
2000, D. stokesi was the eighth most common species and was sighted in 
52.6 percent of 784 surveys (Afzal et al., 2001). Dichocoenia stokesi 
has been observed in low abundance at 17 of 33 monitoring sites in the 
U.S. Virgin Islands and is the 33rd most common species by percent 
cover (Smith, 2013). Off southeast Florida, D. stokesi comprised 6.8 
percent of the coral population between 9 and 32 m depth and was ranked 
the 5th most abundant coral species out of 27 coral species encountered 
(Goldberg, 1973). In surveys of Conch Reef in the Florida Keys in 1995, 
juvenile D. stokesi comprised between approximately two and six percent 
of the overall juvenile coral population, and the highest proportion 
occurred at 14 m and decreased with depth (Edmunds et al., 2004). Off 
South Caicos Island, D. stokesi was most frequently encountered on 
shallow pavement (9 m) and comprised 15 percent of all coral colonies 
counted; however on the deeper spur and groove (18 m) and fore-reef (27 
m), it comprised 2 and 0.7 percent of colonies counted, respectively 
(Steiner, 1999). Bak and Meesters (1999) report that about 50 percent 
of D. stokesi colonies surveyed in Florida and Curacao were in the 10 
to 20 cm size class.
    Between 1996 and 2003, average cover of D. stokesi per habitat type 
ranged from 0.02 to 0.12 percent in the Florida Keys and was highest on 
patch reefs (Somerfield et al., 2008). Of three sites surveyed in 
Bermuda, cover of D. stokesi was 0.02  0.03 percent at one 
site (Dodge et al., 1982). In surveys off Colombia from 1998 to 2004, 
D. stokesi cover ranged from 0.02 to 0.6 percent, but the species was 
only present in nine out of 32 sites (Rodriguez-Ramirez et al., 2010). 
In the Bahamas Archipelago, cover of D. stokesi was on average 0.01 to 
0.02 percent in 2002 to 2004 (Roff et al., 2011). In Dominica, D. 
stokesi was observed in 47 percent of 31 sites surveyed and comprised 
less than one percent cover (Steiner, 2003). Dichocoenia stokesi was 
present on four out of seven fringing reefs off Barbados and comprised 
between 0.1 and 0.6 percent cover (Tomascik and Sander, 1987).
    On remote reefs off southwest Cuba, D. stokesi was observed on 30 
reef front sites at densities of 0.052  0.096 (SD) colonies 
per 10 m transect, but was not observed at any of the 38 surveys of the 
reef crest (Alcolado et al., 2010). In 1,176 sites surveyed in 
southeast Florida and the Florida Keys between 2005 and 2010, density 
of D. stokesi ranged from 0.07 to 2.35 colonies per 10 m\2\ on reef 
zones where they were found, and this species was the eighth most 
abundant species out of 42 coral species encountered (Burman et al., 
2012).
    The SRR and SIR provided the following information on population 
trends of D. stokesi. A comparison of survey data from 19 sites in 
Spaanse Water, Curacao in 1961 and 1992 indicated an 80 percent 
decrease in relative abundance of D. stokesi between the two survey 
periods. In surveys of the Florida Keys between 1995 and 2002 during 
and after a disease outbreak, the average number of D. stokesi colonies 
per 314-m2 site decreased from 44.3 to 11.2, a decline of 
almost 75 percent. The maximum number of D. stokesi colonies per site 
decreased from 95 to 43, and the minimum number of colonies per site 
decreased from ten to one. There was a shift in the size class 
distribution between 1998 and 2002 with a decrease in the frequency of 
smaller size classes and a shift from dominance by smaller size classes 
to a more even distribution across small to larger size classes. Two D. 
stokesi recruits were found after the disease but did not survive to 
the following year. No colonies greater than 25 cm were observed in 
1998, four years later (2002) many colonies greater than 25 cm were 
observed up to 55 cm.
    The public comments did not provide new or supplemental information 
on D. stokesi's population trends, and we did not find any new or 
supplemental information.
    All information on D. stokesi's abundance and population trends can 
be summarized as follows. Dichocoenia stokesi has been characterized as 
usually uncommon but is usually reported as one of the top 10 most 
abundant species where estimates are available. Based on population 
estimates, there are at least tens of millions of D. stokesi colonies 
present in both the Florida Keys and Dry Tortugas. Absolute abundance 
is higher than the estimate from these two locations given the presence 
of this species in many other locations throughout its range. The 
characterization of its occurrence as usually uncommon gives the 
impression of a lower population abundance than is indicated by 
population estimates. Density estimates range from 0.05 to 2.35 
colonies per 10 m\2\. The sometimes low density and small colony size 
result in low percent cover estimates, generally between 0.01 and less 
than 1 percent, and make it difficult to track population trends from 
percent cover data. Trend data indicate D. stokesi has decreased in 
abundance in at least two locations (i.e., the Florida Keys, and a bay 
in Curacao). Presence of juveniles in several locations indicates 
recruitment is occurring. Recovery from severe population declines in 
the Florida Keys after a disease event was not reported seven years 
later. Thus, we conclude that population decline has occurred in some 
locations and that the species' absolute abundance is greater than 
hundreds of millions of colonies.
Other Biological Information
    The SRR and SIR provided the following information on D. stokesi's 
life history. Dichocoenia stokesi is a gonochoric broadcast spawner 
with an overall sex ratio of 2 to 1 (male to female) in southeast 
Florida where a small portion of hermaphroditic colonies (approximately 
18 percent) were observed. Minimum size at reproduction was 160 cm\2\, 
and two potential spawning events per year were

[[Page 53931]]

inferred: one in late August/early September and a second in October. 
Recruitment levels, inferred from the presence of juveniles, is 
intermediate compared to other Caribbean coral species. Very low 
densities of Dichocoenia juveniles (approximately one percent of total 
juvenile colonies) have been observed in the Netherlands Antilles. Mean 
D. stokesi juvenile density among 566 sites surveyed during 1999 to 
2009 averaged 0.11 per m\2\ but reached as high as one juvenile per 
m\2\ in certain habitats. The annual growth rate of D. stokesi has been 
reported as 2 to 7 mm per year in diameter and 2 to 5.2 mm per year in 
height.
    The public comments did not provide new or supplemental information 
on the life history of D. stokesi. Supplemental information we found on 
the life history of D. stokesi includes the following. Chiappone and 
Sullivan (1996) reported density of juvenile D. stokesi range from 0.02 
to 0.26 per m\2\ at five out of nine sites surveyed in the Florida Keys 
between 1993 and 1994. Darling et al. (2012) performed a biological 
trait-based analysis to categorize coral species into four life history 
strategies: Generalist, weedy, competitive, and stress-tolerant. The 
classifications were primarily separated by colony morphology, growth 
rate, and reproductive mode. Dichocoenia stokesi was classified as a 
``stress-tolerant'' species, thus likely more tolerant of environmental 
stress.
    The SRR and SIR provided the following other biological information 
about D. stokesi. The mounding morphology and large corallite diameter 
of D. stokesi enhance turbulence near the surface of colonies. This 
should, in turn, enhance mass transfer, which affects photosynthesis 
and respiration in D. stokesi as well as prey capture and nutrient 
uptake. Thresholds for uptake of inorganic nitrogen in D. stokesi have 
been reported to be fairly low (150 nM), giving it a potential 
advantage in nutrient-poor conditions.
    The public comments did not provide new or supplemental information 
on D. stokesi's biology. Supplemental information we found on D. 
stokesi's biology includes the following. At 76 sites surveyed in the 
Florida Keys during the 2010 cold-water event, approximately 15 percent 
of D. stokesi paled, and approximately one percent bleached. Mortality 
was approximately four percent (The Nature Conservancy, 2010).
Susceptibility to Threats
    The threat susceptibility information from the SRR and SIR was 
interpreted in the proposed rule for D. stokesi's vulnerabilities to 
threats as follows: High vulnerability to disease; moderate 
vulnerability to ocean warming, acidification, trophic effects of 
fishing, and sedimentation; and low vulnerability to sea level rise, 
predation, and collection and trade.
    The SRR and SIR provided the following information on the 
susceptibility of D. stokesi to ocean warming. Of the 28 coral species 
that bleached along the Florida reef tract from Martin County through 
the lower Florida Keys from 2005 to 2007, D. stokesi had the lowest 
bleaching prevalence. During the 2005 Caribbean mass-bleaching event, 
it ranked 16th of 21 species in bleaching prevalence in Barbados and 
was observed to be bleaching-tolerant in the U.S. Virgin Islands.
    Van Woesik et al. (2012) developed a coral resiliency index based 
on biological traits and processes to evaluate extinction risk due to 
bleaching. Evaluations were performed at the genus level. They rated 
the resiliency of Dichocoenia as 0 out of a range of -6 to 7 observed 
in other coral genera. Less than or equal to -3 was considered highly 
vulnerable to extinction, and greater than or equal to 4 was considered 
highly tolerant. Thus, Dichocoenia was rated in the middle.
    The public comments did not provide new or supplemental information 
on the susceptibility of D. stokesi to ocean warming. Supplemental 
information we found on the susceptibility of D. stokesi to ocean 
warming includes the following. During the 1998 bleaching event, an 
average of 20 percent of D. stokesi colonies were greater than 50 
percent bleached in the lower Florida Keys and Dry Tortugas; however, 
this was the lowest of 14 species that bleached (Waddell, 2005). Of the 
22 species monitored off Roatan, Honduras, D. stokesi was one of eight 
species that did not bleach during the 1998 bleaching event (Riegl et 
al., 2009).
    During the 2005 temperature anomaly, D. stokesi colonies were fully 
bleached around La Parguera, Puerto Rico but were less frequently 
bleached at other locations around Puerto Rico (Waddell and Clarke, 
2008). Off of Mona and Desecheo Islands, Puerto Rico, about 25 percent 
of D. stokesi paled and about 10 percent bleached; in the 16 coral 
species surveyed, bleaching ranged from less than five percent to 
approximately 60 percent of colonies (Bruckner and Hill, 2009). During 
the 2005 bleaching event, approximately 30 percent of D. stokesi 
colonies on six reefs bleached in Barbados, and D. stokesi around Grand 
Cayman experienced total bleaching (Wilkinson and Souter, 2008). None 
of the monitored D. stokesi colonies in the U.S. Virgin Islands 
bleached, and 67 percent paled during the 2005 bleaching event (Smith 
et al., 2013b). In the Florida Keys, D. stokesi ranked 19th out of 25 
species in amount of mortality during the 2005 bleaching event (Lirman 
et al., 2011).
    All sources of information are used to describe D. stokesi's 
susceptibility to ocean warming as follows. Reported bleaching of D. 
stokesi ranges from zero to about 60 percent. While reported bleaching 
of D. stokesi is temporally and spatially variable, compared to other 
Caribbean coral species, D. stokesi appears to be among the less 
susceptible to temperature-induced bleaching. Additionally, a report 
from the Florida Keys indicates that bleaching-induced mortality of D. 
stokesi was among the lowest compared to other Caribbean coral species. 
Thus, we conclude that D. stokesi has some susceptibility to ocean 
warming. However, the available information does not support a more 
precise description of susceptibility.
    The SRR and SIR provided the following information on the 
susceptibility of D. stokesi to acidification. No specific research has 
addressed the effects of acidification on the genus Dichocoenia. 
However, most corals studied have shown negative relationships between 
acidification and growth, and acidification is likely to contribute to 
reef destruction in the future. While ocean acidification has not been 
demonstrated to have caused appreciable declines in coral populations 
so far, it is considered a significant threat to corals by 2100.
    The public comments did not provide new or supplemental information 
on the susceptibility of D. stokesi to acidification, and we did not 
find any new or supplemental information.
    All sources of information are used to describe D. stokesi's 
susceptibility to acidification as follows. There is uncertainty about 
how D. stokesi will respond to ocean acidification, but based on the 
negative effects of acidification on growth of most corals, D. stokesi 
likely has some susceptibility to acidification. The available 
information does not support a more precise description of 
susceptibility.
    The SRR and SIR provided the following information on D. stokesi's 
susceptibility to disease. Black band disease, dark spot syndrome, and 
white plague have been reported to affect D. stokesi. In an outbreak of 
white plague in St. Lucia in 1997, six surveyed colonies of D. stokesi 
were infected, and average tissue mortality was about 65 percent. In 
surveys in Dominica

[[Page 53932]]

between 2000 and 2002, D. stokesi was one of four coral species most 
commonly affected by disease, and white plague predominantly affected 
larger-sized colonies. Of 17 species affected by white plague in the 
Florida Keys, D. stokesi was the most susceptible.
    The public comments did not provide new or supplemental information 
on the susceptibility of D. stokesi to disease. Supplemental 
information we found on the susceptibility of D. stokesi to disease 
includes the following. In 1991, an outbreak of white plague was 
observed on Mona Island, Puerto Rico that affected 14 species, with the 
highest prevalence among small, massive corals including D. stokesi, 
many of which died within one to two weeks (Waddell, 2005). In Mexico, 
disease was prevalent on approximately one percent of D. stokesi 
colonies surveyed in 2004 (Ward et al., 2006).
    During an outbreak of white plague type II in the Florida Keys in 
1995, mortality of D. stokesi averaged 26 percent and ranged from 0 to 
38 percent (Richardson et al., 1998). The disease routinely caused 
whole colony mortality within two to three days due to its infection of 
small coral colonies (usually less than 10 cm in diameter) and 
aggressive progression rate (up to 2 cm per day; Richardson, 1998). 
Between 1996 and 1998, out of 160 monitoring stations at 40 sites in 
the Florida Keys, the number of stations with D. stokesi colonies 
affected by disease increased through time with two stations affected 
in 1996, 22 in 1997, and 45 in 1998 (Porter et al., 2001). However, no 
white plague was observed in D. stokesi in 2002 at the sites with the 
reported outbreak in 1995 (Richardson and Voss, 2005).
    Disease surveys at St. Croix, U.S. Virgin Islands during the summer 
of 2001 revealed that D. stokesi had the highest prevalence of white 
plague type II out of seven species infected and the highest disease-
related mortality (Kaczmarsky et al., 2005). The prevalence of white 
plague type II on D. stokesi was 41 percent at one location and 60 
percent at a second site. Of 107 D. stokesi colonies, 38 were infected, 
and 26 percent of the infected colonies, or 9.4 percent of the sample 
population, died within two months (Kaczmarsky et al., 2005). After the 
2005 bleaching event, 100 percent of monitored D. stokesi colonies in 
the U.S. Virgin Islands were infected with disease in 2006, but none of 
the colonies experienced total colony mortality (Smith et al., 2013b).
    All sources of information are used to describe D. stokesi's 
susceptibility to disease as follows. Although D. stokesi is 
susceptible to several diseases, the most severe impacts have been the 
result of white plague. Low prevalence of diseased D. stokesi colonies 
have been reported from some locations, but outbreaks of white plague 
have caused rapid and substantial mortality in some other sites. 
Outbreaks in Puerto Rico and St. Lucia, while affecting D. stokesi, do 
not appear to have caused as severe mortality as in the Florida Keys 
and U.S. Virgin Islands. Thus, we conclude that D. stokesi has high 
susceptibility to disease.
    The SIR and SRR did not provide any species-specific information on 
the trophic effects of fishing on D. stokesi. The public comments did 
not provide any new or supplemental information on the trophic effects 
of fishing on D. stokesi, and we did not find any new or supplemental 
information. However, due to the level of reef fishing conducted in the 
Caribbean, coupled with Diadema die-off and lack of significant 
recovery, competition with algae can adversely affect coral 
recruitment. Based on D. stokesi's inferred recruitment rates, we 
conclude that it likely has low susceptibility to trophic effects of 
fishing.
    The SRR and SIR provided the following information on 
susceptibility of D. stokesi to sedimentation. A laboratory study 
examining oil/sediment rejection indicated that out of 19 Caribbean 
coral species examined, D. stokesi was intermediate in the rate of 
sediment removal from its tissues. In laboratory experiments, D. 
stokesi exhibited significant increases in respiration after 3 days of 
exposure to turbidity levels of 28 to 30 NTU, which are within 
allowable levels as regulated by the State of Florida for coastal 
construction projects. While light levels and photosynthesis were not 
affected, after six days of exposure to 14 to 16 NTU of turbidity, 
gross photosynthesis to respiration ratios were less than one in this 
species, and excessive mucus production was observed.
    The public comments did not provide new or supplemental information 
on the susceptibility of D. stokesi to sedimentation. Supplemental 
information we found on the susceptibility of D. stokesi to 
sedimentation includes the following. The large calices, number of 
septa, and calical relief of D. stokesi give this species the 
capability to remove both fine sediment and larger grain sizes through 
polyp distension (Hubbard and Pocock, 1972).
    All sources of information are used to describe D. stokesi's 
susceptibility to sedimentation as follows. Dichocoenia stokesi is more 
tolerant of sedimentation than other coral species as it has the 
ability to remove both larger grain size and finer sediment. However, 
prolonged exposure (several days) to turbidity has been shown to cause 
physiological stress. We conclude that D. stokesi has some 
susceptibility to sedimentation. However, the available information 
does not support a more precise description of susceptibility.
    The SRR and SIR did not provide any species or genus information on 
the susceptibility of D. stokesi to nutrients but provided the 
following. Land-based sources of pollution (including nutrients) often 
act in concert rather than individually and are influenced by other 
biological (e.g., herbivory) and hydrological factors. Collectively, 
land-based sources of pollution are unlikely to produce extinction at a 
global scale; however, they may pose significant threats at local 
scales and reduce the resilience of corals to bleaching.
    The public comments did not provide new or supplemental information 
on the susceptibility of D. stokesi to nutrients, and we did not find 
any new or supplemental information. Based on our knowledge that 
nutrients in general have a negative effect on corals, we conclude that 
D. stokesi has some level of susceptibility to nutrients, but the 
available information does not support a more precise description of 
susceptibility.
    The SRR and SIR provided the following information on the 
susceptibility of D. stokesi to predation. Dichocoenia stokesi is 
minimally affected by predation. Sponges such as Chondrilla nucula and 
Ectoplaysia ferox can overgrow and cause tissue loss in D. stokesi, 
especially if unchecked by spongivores. Dichocoenia stokesi had the 
highest density of boring bivalves (average 7.5 bivalves per colony) of 
the three coral species examined.
    The public comments provided supplemental information on D. 
stokesi's susceptibility to predation. Predation by Coralliophila 
snails was recorded on 1.8 percent of the 502 D. stokesi colonies 
assessed for condition in 2012 surveys in the Florida Keys (Miller et 
al., 2013). We did not find any new or supplemental information on the 
susceptibility of D. stokesi to predation.
    All sources of information confirm that predation does not appear 
to significantly affect D. stokesi. Thus, we conclude that D. stokesi 
has low susceptibility to predation.
    The SRR and SIR provided the following information on the 
susceptibility of D. stokesi to collection and trade. Collection and 
trade are not considered a threat to D. stokesi. The

[[Page 53933]]

public comments did not provide new or supplemental information. 
Supplemental information we found on collection and trade includes the 
following. Collection and trade of D. stokesi appear to be low and 
primarily for scientific purposes. Gross exports between 2000 and 2012 
averaged 35 corals per year (data available at http://trade.cites.org). 
Thus, we conclude that D. stokesi has low susceptibility to collection 
and trade.
    The SRR and SIR did not provide species-specific information on the 
effects of sea level rise on D. stokesi. The SRR described sea level 
rise as an overall low to medium threat for all coral species. The 
public comments did not provide new or supplemental information on D. 
stokesi's susceptibility to sea level rise, and we did not find any new 
or supplemental information. Thus, we conclude that D. stokesi has some 
susceptibility to sea level rise, but the available information does 
not provide a more precise description of susceptibility.
Regulatory Mechanisms
    In the proposed rule, we relied on information from the Final 
Management Report for evaluating the existing regulatory mechanisms for 
controlling threats to all corals. However, we did not provide any 
species-specific information on the regulatory mechanism or 
conservation efforts for D. stokesi. Public comments were critical of 
that approach, and we therefore attempt to analyze regulatory 
mechanisms and conservation efforts on a species basis, where possible, 
in this final rule. Records confirm that Dichocoenia stokesi occurs in 
nine Atlantic ecoregions that encompass 26 kingdom's and countries' 
EEZs. The 26 kingdoms and countries are Antigua & Barbuda, Bahamas, 
Barbados, Belize, Colombia, Costa Rica, Cuba, Dominica, Dominican 
Republic, French Antilles, Grenada, Guatemala, Haiti, Kingdom of the 
Netherlands, Honduras, Jamaica, Mexico, Nicaragua, Panama, St. Kitts & 
Nevis, St. Lucia, St. Vincent & Grenadines, Trinidad and Tobago, United 
Kingdom (British Overseas Territories), United States (including U.S. 
Caribbean Territories), and Venezuela. The regulatory mechanisms 
relevant to D. stokesi, described first as a percentage of the above 
countries and kingdoms that utilize them to any degree, and second as 
the percentages of those countries and kingdoms whose regulatory 
mechanisms may be limited in scope, are as follows: General coral 
protection (31 percent with 12 percent limited in scope), coral 
collection (50 percent with 27 percent limited in scope), pollution 
control (31 percent with 15 percent limited in scope), fishing 
regulations on reefs (73 percent with 50 percent limited in scope), 
managing areas for protection and conservation (88 percent with 31 
percent limited in scope). The most common regulatory mechanisms in 
place for D. stokesi are reef-fish fishing regulations and area 
management for protection and conservation. However, half of the reef-
fish fishing regulations are limited in scope and may not provide 
substantial protection for the species. General coral protection and 
collection laws, along with pollution control laws, are much less 
common regulatory mechanisms for the management of D. stokesi.
Vulnerability to Extinction
    As explained above in the Risk Analyses section, a species' 
vulnerability to extinction results from the combination of its spatial 
and demographic traits, threat susceptibilities, and consideration of 
the baseline environment and future projections of threats. The SRR 
stated that the factors that increase the potential extinction risk for 
D. stokesi include documented population-level impacts from disease. 
Factors that reduce potential extinction risk are relatively high 
abundance and persistence across many habitat types, including 
nearshore and mesophotic reefs. Residency in a wide range of habitat 
types suggests the species has a wide tolerance to environmental 
conditions and, therefore, better capacity to deal with changing 
environmental regimes.
    Subsequent to the proposed rule, we received and gathered 
supplemental species- or genus-specific information, described above, 
that expands our knowledge regarding the species' abundance, 
distribution, and threat susceptibilities. We developed our assessment 
of the species' vulnerability to extinction using all the available 
information. As explained in the Risk Analyses section, our assessment 
in this final rule emphasizes the ability of the species' spatial and 
demographic traits to moderate or exacerbate its vulnerability to 
extinction, as opposed to the approach we used in the proposed rule, 
which emphasized the species' susceptibility to threats.
    The following characteristics of D. stokesi, in conjunction with 
the information described in the Corals and Coral Reefs section, Coral 
Habitat sub-section, and Threats Evaluation section above, affect its 
vulnerability to extinction currently and over the foreseeable future. 
Although it is geographically located in the heavily disturbed 
Caribbean, D. stokesi occurs in a wide range of habitats, including 
mesophotic reefs, back- and fore-reef environments, rocky reefs, 
lagoons, spur-and-groove formations, channels, and occasionally at the 
base of reefs. This distribution in a wide range of environments 
suggests the species will be better able to withstand changing 
environmental conditions and moderates vulnerability to extinction over 
the foreseeable future because the numerous types of reef environments 
in which the species occurs are predicted, on local and regional 
scales, to experience highly variable thermal regimes and ocean 
chemistry at any given point in time. It has been reported in water 
depths ranging from 2 to 72 m. Deeper areas of D. stokesi's range will 
usually have lower temperatures than surface waters, and acidification 
is generally predicted to accelerate most in waters that are deeper and 
cooler than those in which the species occurs. The species is highly 
susceptible to disease, and outbreaks have resulted in high colony 
mortality in some locations in its range. However, D. stokesi's 
abundance has been estimated as at least tens of millions of colonies 
in both the Florida Keys and Dry Tortugas and is higher than the 
estimate from these two locations due to the occurrence of the species 
in many other areas throughout its range. Additionally, sexual 
recruitment, as evidenced by presence of juvenile colonies, is 
comparatively higher than many other Caribbean coral species, enhancing 
recovery potential from mortality events, thus moderating vulnerability 
to extinction. The combination of wide habitat occupancy, abundance, 
life history characteristics, and depth distribution, combined with 
spatial variability in ocean warming and acidification across the 
species' range, moderates vulnerability to extinction because the 
increasingly severe conditions expected in the foreseeable future will 
be non-uniform, and there will likely be a large number of colonies 
that are either not exposed or do not negatively respond to a threat at 
any given point in time.
Listing Determination
    In the proposed rule, using the determination tool formula 
approach, D. stokesi was proposed for listing as threatened because of: 
High vulnerability to disease (C); moderate vulnerability to ocean 
warming (E) and acidification (E); moderate overall distribution (based 
on narrow geographic distribution and wide depth distribution (E); 
restriction to the

[[Page 53934]]

Caribbean (E); and inadequacy of regulatory mechanisms (D).
    In this final rule, we changed the listing determination for D. 
stokesi from threatened to not warranted. We made this determination 
based on a more species-specific and holistic assessment of whether 
this species meets the definition of either a threatened or endangered 
coral, including more appropriate consideration of the buffering 
capacity of this species' spatial and demographic traits to lessen its 
vulnerability to threats. Thus, based on the best available information 
above on D. stokesi's spatial structure, demography, threat 
susceptibilities, and management, none of the five ESA listing factors, 
alone or in combination, are causing this species to be likely to 
become endangered throughout its range within the foreseeable future, 
and thus it is not warranted for listing at this time because:
    (1) Dichocoenia stokesi's distribution in depths of two to 72 m in 
heterogeneous habitats, including mesophotic reefs, back- and fore-reef 
environments, rocky reefs, lagoons, spur-and-groove formations, 
channels, and occasionally at the base of reefs, throughout the 
Caribbean basin reduces exposure to any given threat event or adverse 
condition that does not occur uniformly throughout the species' range. 
As explained above in the Threats Evaluation section, we have not 
identified any threat that is expected to occur uniformly throughout 
the species range within the foreseeable future;
    (2) Dichocoenia stokesi is usually reported in the top ten most 
abundant coral species in the Caribbean, and its total absolute 
abundance is at least tens of millions of colonies based on estimates 
from two locations. Absolute abundance is higher than estimates from 
these locations since it occurs in many other locations throughout its 
range. This provides buffering capacity in the form of absolute numbers 
of colonies and variation in susceptibility between individual 
colonies. As discussed in the Corals and Coral Reefs section above, the 
more colonies a species has, the lower the proportion of colonies that 
are likely to be exposed to a particular threat at a particular time, 
and all individuals that are exposed will not have the same response;
    (3) Dichocoenia stokesi occurs in most reef habitats, including 
mesophotic reefs, back- and fore-reef environments, rocky reefs, 
lagoons, spur-and-groove formations, channels, and occasionally at the 
base of reefs, indicating wide tolerance of environmental conditions 
and better capacity to deal with changing environmental regimes; and
    (4) Presence of juvenile D. stokesi colonies indicates that 
recruitment is likely occurring, enhancing recovery potential from 
mortality events.
    Notwithstanding the projections through 2100 that indicate 
increased severity over time of the three high importance threats, the 
combination of these biological and environmental characteristics 
indicates that the species possesses sufficient buffering capacity to 
avoid being in danger of extinction within the foreseeable future 
throughout its range. This species' extinction risk may increase in the 
future if global threats continue and worsen in severity and the 
species' exposure to the threats increases throughout its range. Should 
the species experience reduced abundance or range constriction of a 
certain magnitude, the ability of these characteristics to moderate 
exposure to threats will diminish. However, D. stokesi is not likely to 
become of such low abundance or so spatially fragmented as to be in 
danger of extinction due to depensatory processes, the potential 
effects of environmental stochasticity, or the potential for mortality 
from catastrophic events within the foreseeable future throughout its 
range. Therefore, D. stokesi is not warranted for listing at this time 
under any of the listing factors, and we withdraw our proposal to list 
the species as threatened.

Genus Orbicella (formerly Montastraea)

Introduction
    The SRR and SIR provided the following information on the taxonomy 
and morphology of the genus Montastraea. The genus Montastraea 
contained four Caribbean species: M. cavernosa, M. annularis, M. 
faveolata, and M. franksi. Prior to the 1990s, M. annularis, M. 
faveolata, and M. franksi were considered one species, M. annularis. 
However, M. annularis was broken into the three sibling species based 
on behavioral, biochemical, and morphological criteria. These three 
species are often grouped into the M. annularis species complex. 
Subsequent reproductive and genetic studies have generally supported 
the partitioning of the complex into three species. Montastraea 
faveolata is the most genetically distinct while M. annularis and M. 
franksi are less so.
    The public comments provided the following new information on 
Montastraea's taxonomy. In 2012, the genus Montastraea was split, and 
M. annularis, M. faveolata, and M. franksi were assigned to the genus 
Orbicella (Budd et al., 2012). From this point forward, we will refer 
to the genus and species by their current taxonomic classification in 
the genus Orbicella. We did not find any new or supplemental 
information on Orbicella's taxonomy or morphology.
    Some studies report on the species complex rather than individual 
species since visual distinction can be difficult from video or 
photographic surveys or in small colonies where morphology is more 
difficult to discern. This section will report information on the 
species complex and on O. annularis from studies pre-dating 1994 when 
the species was split into three nominal species.
Spatial Information
    The SRR and SIR provide the following information on Orbicella's 
distribution, habitat, and depth range. The species complex has been 
found at depths to 90 m. It is dominant on mesophotic reefs in Puerto 
Rico and the U.S. Virgin Islands at depths of 30 to 45 m, and it is 
found at depths up to 70 to 90 m in these locations.
    The public comments did not provide new or supplemental information 
on Orbicella's distribution, habitat, or depth range. Supplemental 
information we found on Orbicella's depth range includes the following. 
All three species occupy a large depth range. Although there is depth 
overlap in species occurrence, there is larger variance and overlap in 
species abundances in shallow versus deep water (Pandolfi and Budd, 
2008). Orbicella faveolata tends to have the shallowest depth 
distribution, and O. franksi tends to have the deepest (Pandolfi and 
Budd, 2008; Weil and Knowlton, 1994). At three study sites in Belize, 
O. faveolata was the most abundant member of the species complex 
between 2 and 5 m depth; O. annularis was the most abundant at depths 
of 10 to 15 m, and O. franksi was the most abundant at depths of 20 to 
30 m (Pandolfi and Budd, 2008). Orbicella annularis species complex can 
be relatively abundant at mesophotic depths in the Bahamas, Belize, 
Jamaica, Puerto Rico, U.S. Virgin Islands, and Curacao (Kahng et al., 
2010).
Demographic Information
    The SRR and SIR provided the following information on abundance and 
population trends of the Orbicella annularis species complex. The 
species complex has historically been a dominant component on Caribbean 
coral reefs, characterizing the so-called ``buttress zone'' and 
``annularis zone'' in the classical descriptions of Caribbean reefs. 
The species complex is the major reef-builder in the greater Caribbean,

[[Page 53935]]

since the die-off of Acropora spp., due to their large size and high 
abundance.
    Numerous examples of population decline of the Orbicella annularis 
species complex were described, and the results are summarized as 
follows. Decline in the Florida Keys between the late 1970s and 2003 
was approximately 80 to 95 percent, with further losses during the 2012 
cold weather event. Decadal-scale declines across the remote islands of 
Navassa, Mona, and Desecheo in the central Caribbean impacted 85 
percent of colonies found there. In the U.S. Caribbean (U.S. Virgin 
Islands and Puerto Rico), an 80 to 90 percent decline has been reported 
over the past two decades. Percent cover was reportedly stable in 
Curacao in the mid-1970s, an 85 percent increase in partial mortality 
occurred between 1998 and 2005. Between 1975 and 1998 at Glovers Reef 
in Belize, a 38 to 75 percent decline in relative cover occurred with a 
further 40 percent decline since. Colonies in Colombia were stable 
between 1998 and 2003 although demographic changes imply some degree of 
decline. Surveys of population structure across five countries found a 
significant increase in small ramets (tissue isolates that are 
genetically identical but physiologically separate from the parent 
colony) less than 500 cm\2\ (211 percent for O. annularis, 168 percent 
for O. faveolata, 137 percent for O. franksi), while the proportion of 
large (1,500- 30,000 cm\2\), completely live colonies declined by 51 to 
57 percent.
    The public comments did not provide new or supplemental information 
on Orbicella's abundance and population trends. Supplemental 
information we found on Orbicella's abundance and population trends is 
provided as follows. In a survey of 185 sites in five countries 
(Bahamas, Bonaire, Cayman Islands, Puerto Rico, and St. Kitts and 
Nevis) between 2010 and 2011, Orbicella annularis species complex 
exhibited mean tissue mortality of 29 to 66 percent, which was higher 
than other species exhibiting mean 8 to 17 percent tissue mortality. 
Total mortality of O. annularis species complex were observed (five to 
seven percent of the total); however mortality of large colonies mostly 
resulted in multiple smaller ramets Mortality was attributed primarily 
to outbreaks of white plague and yellow band disease, which emerged as 
corals began recovering from mass bleaching events. This was followed 
by increased predation and removal of live tissue by damselfish to 
cultivate algal lawns (Bruckner, 2012a).
    In 1998 O. annularis species complex covered more of the benthos 
than any other coral taxon at nine monitored sites off Mona and 
Desecheo Islands, Puerto Rico: 47 percent on reefs off Desecheo Island 
and 32 percent off Mona Island. In 2008 live cover of O. annularis 
species complex ranged from 0 to 14 percent with 95 percent decline off 
Desecheo Island and 78 percent decline off Mona Island. This was 
accompanied with large changes in the size frequency distribution and 
extent of partial mortality, with size structure remaining constant. 
The amount of living tissue declined by 55 percent due to partial 
mortality affecting medium and large colonies, with an increase in the 
number of colonies with small (less than 10 cm diameter) tissue 
remnants. Sponges and macroalgae colonized newly exposed area, and 
sponges appeared to be preventing re-sheeting of tissue remnants. No 
Orbicella spp. recruits were observed during the ten year study 
(Bruckner and Hill, 2009).
    Surveys at three reefs in western Curacao in 1998 found 46 percent 
of all corals were O. annularis species complex. In 2005, O. annularis 
species complex remained the dominant coral species but declined in 
abundance to 38 percent of the overall coral population (decreases in 
abundance occurred in O. faveolata and O. annularis, but not O. 
franksi). In 1998 mean diameter of O. annularis species complex 
colonies were 62 cm and less than 10 percent of all O. annularis 
species complex colonies were less than 30 cm in diameter. Partial 
mortality of O. annularis species complex increased 85 percent between 
1997 and 2005 with losses of O. annularis and O. faveolata (partial 
mortality 42 to 48 percent and total mortality 6 percent for the two 
species combined) larger than O. franksi. The most significant losses 
were due to yellow band disease and white plague. No recruits of O. 
annularis species complex were observed between 1997 and 2005 in 
transects or on skeletons of tagged colonies exposed through mortality 
from disease (Bruckner and Bruckner, 2006a).
    McClanahan and Muthiga (1998) surveyed 20 patch reefs in Glovers 
Reef atoll off Belize between 1996 and 1997 and compared their results 
to surveys of 16 patch reefs in the same general area conducted between 
1970 and 1971. They found that O. annularis species complex experienced 
an overall 62 percent decrease in cover. Average cover of O. annularis 
species complex was seven percent in 1996 and 1997.
    The O. annularis species complex often makes up the largest 
proportion of coral cover on Caribbean reefs. In surveys conducted on 
four reefs in Biscayne National Park, Florida in 1981, cover of O. 
annularis species complex ranged between approximately 25 and 50 
percent on three of the reefs, and no O. annularis species complex 
colonies were observed in transects on the fourth reef (Burns, 1985). 
In stratified random surveys in 2007-2008, O. annularis species complex 
was the dominant coral by percent cover in the Red Hind Marine 
Conservation District off St. Thomas, U.S. Virgin Islands, at depths of 
34 to 47 m. Orbicella annularis species complex averaged 15 percent 
cover (range zero to 48 percent) and made up 92 percent of the 25 
percent average coral cover (Nemeth et al., 2008).
    In a survey of 185 sites in five countries (Bahamas, Bonaire, 
Cayman Islands, Puerto Rico, and St. Kitts and Nevis) in 2010 to 2011, 
density of O. annularis species complex ranged from 0.3 to 2.7 colonies 
per m\2\ and comprised between 9 and 30 percent of all corals greater 
than 4 cm diameter. The mean diameter ranged from 44 to 89 cm, and the 
size structure (planar surface area) had a bell shaped distribution, 
with only a few colonies less than 500 cm\2\ or greater than 10,000 
cm\2\ (Bruckner, 2012a).
    In surveys of juvenile corals (less than 4 cm diameter) on nine 
reefs in the Florida Keys between 1993 and 1994, density of O. 
annularis species complex ranged between 0.02 and 0.04 juvenile corals 
per m\2\ on six of the nine reefs. Density of O. annularis species 
complex juveniles was correlated with non-juvenile O. annularis species 
complex density and with depth. The majority of non-juveniles were 
smaller than the reproductive size of 100 cm\2\ (Chiappone and 
Sullivan, 1996).
    Surveys in Bonaire in 2008 showed that the O. annularis species 
complex dominated coral cover in depths less than 20 m and cover was 
similar to that reported in 1982. However, all sites surveyed in 2008 
showed signs of disease and partial mortality in a large number of the 
massive colonies, and many were reduced to a patchwork of live tissue 
and dead areas colonized by algae (Stokes et al., 2010).
    At 25 sites surveyed in Bonaire in 2011, O. annularis species 
complex was the dominant coral taxa occupying approximately 20 to 25 
percent of the benthos and making up 46 percent of the total live coral 
cover. It was dominant in terms of abundance, making up approximately 
27 percent of all corals. Orbicella annularis was significantly more 
abundant than O. franksi and O. faveolata on the northern reefs but not 
on southern reefs. Most colonies were between 30 and 80 cm diameter 
with size structure of O. annularis species complex in a bell

[[Page 53936]]

shaped distribution around this range; there were few colonies less 
than 20 cm and few very large colonies greater than 200 cm, with a 
small peak at the 150 to 199 cm range. There was a notable absence of 
colonies less than 10 cm diameter (as measured by the skeleton, not 
live tissue) and an absence of recruits. A total of 73 out of 1602 
colonies (4.5 percent) had completely died. Surviving colonies (n=1529) 
had a mean of 28 percent partial mortality. On average, each colony was 
divided into 6.6 tissue remnants. Several sites contained a high 
abundance of large, unblemished O. annularis species complex colonies 
(Bruckner, 2012c).
    Between 1999 and 2009, overall cover of O. annularis species 
complex in the Florida Keys declined, but differed by habitat type 
(Ruzicka et al., 2013). Percent cover declined on the deep and shallow 
fore-reefs but remained stable on patch reefs (Ruzicka et al., 2013). 
The 2010 cold-water event reduced cover of O. annularis species complex 
from 4.4 percent to 0.6 percent on four patch reefs in the upper and 
middle Florida Keys. Greater than 50 percent of O. annularis species 
complex colonies across all size classes suffered lethal or severe 
mortality, and 93 percent of all O. annularis species complex colonies 
surveyed suffered complete or partial mortality. The species complex 
suffered the highest mortality of all coral species affected (Colella 
et al., 2012). A comparison of 1995 and 2005 surveys of O. annularis 
species complex at 13 patch reefs in the Florida Keys reported ten 
sites had between 5 and 40 percent more dead areas (Gischler, 2007).
    Density of juvenile O. annularis species complex increased from 
0.07 juveniles per m\2\ prior to 2008, to 0.15 juveniles per m\2\ and 
continued at 0.12 juveniles per m\2\ in 2009 at 4 km area on the south 
side of St. John, U.S. Virgin Islands that has been monitored for 16 
years. These densities were driven by seven to nine colonies per year, 
and the increased density did not extend outside the initial survey 
area when expanded to other areas around St. John. While not possible 
to distinguish the species in the field, the authors conclude juveniles 
were most likely O. annularis due to the abundance of O. annularis on 
adjacent reefs and the rarity of the presence of the other two species 
in water less than 9 m (Edmunds et al., 2011).
    At Yawzi Point, St. John, U.S. Virgin Islands, the percentage of 
total coral cover declined by more than 50 percent between 1987 to 
1998, from 45 percent to 20 percent. In 1988, 94 percent of the coral 
cover at Yawzi was O. annularis species complex mostly O. annularis (97 
percent), with a few colonies of O. faveolata (6 percent). Despite a 
reduction in total cover, O. annularis species complex remained 
spatially dominant in 1998 at 96 percent of the coral cover (Edmunds, 
2002). Coral cover at this site again declined an additional 65 percent 
between 1999 and 2011 to seven percent cover, with O. annularis species 
complex remaining dominant at 77 percent of the coral cover (Edmunds, 
2013).
    At Tektite Reef, St. John, U.S. Virgin Islands, total coral cover 
increased from 32 percent in 1987 to 43 percent in 1998 but then 
decreased to 29 percent in 2011 (Edmunds, 2002; Edmunds, 2013). In 
1988, 79 percent of the species complex was O. annularis, with lesser 
amounts of O. faveolata (one percent) and O. franksi (21 percent) 
(Edmunds, 2002). Greater than 72 percent of coral was O. annularis 
species complex in all survey years (Edmunds, 2013).
    Surveys of the Flower Garden Banks between 1974 and 1980 found 
cover of O. annularis species complex between approximately 23 and 40 
percent in areas less than 36 m depth (Bright et al., 1984). Orbicella 
annularis species complex was the dominant coral between 2002 and 2003 
at 32 percent cover (Aronson et al., 2005). In random surveys between 
2002 and 2006, O. annularis species complex (predominantly O. franksi) 
was the dominant coral in the Flower Garden Banks comprising between 27 
and 40 percent benthic cover (Hickerson et al., 2008). In permanent 
photo quadrats (8 m\2\ total), cover of O. annularis species complex 
(as measured by planar surface area of individual colonies) fluctuated 
between approximately 20 and 45 percent cover in the East Flower 
Gardens between 1992 and 2006 with periods of sharp increase and 
decrease in cover (Hickerson et al., 2008). Cover in west Flower 
Gardens was between 22 and 40 percent over the same time period and had 
less annual variability and a generally increasing or stable trend 
through time (Hickerson et al., 2008).
    Surveys of five sites in the Mexican Yucatan in 1985 and 2005 
revealed a decrease in relative cover of O. annularis species complex. 
At four out of the five sites, cover of O. annularis species complex 
decreased from between approximately 50 and 60 percent in 1985 to 
between approximately 10 and 25 percent in 2005. The fifth site had a 
less dramatic decrease in relative cover from approximately 35 percent 
to 30 percent cover during this 20-year interval. Disease appeared to 
be the main cause of decline, but hurricanes may have also played a 
role (Harvell et al., 2007).
    Size transition matrices were derived from Orbicella growth, 
mortality, and recruitment rates between 1998 and 2003 from four sites 
in the lower Florida Keys. Forecasting 15 years into the future 
predicted a steady decline in all size classes except the smallest 
(less than 5 cm) due to insufficient recruitment to offset mortality 
and low growth rates of the smaller size classes. Mortality rates were 
assumed at approximately 40 percent for the smallest size class 
declining to about 5 percent for the largest (Smith and Aronson, 2006).
    All information on Orbicella's abundance and population trends can 
be summarized as follows. The O. annularis species complex historically 
dominated fore-reef sites throughout the Caribbean both in abundance 
and cover and formed dense assemblages of large, hundreds-of-years old 
colonies and few small colonies (Bruckner, 2012a). However, recent 
declines in O. annularis species complex cover have been reported. 
Major declines range from approximately 50 to 95 percent in locations 
including Puerto Rico, Belize, the Florida Keys, Mexico, and the U.S. 
Virgin Islands, and lower levels of decline (5 to 33 percent) have been 
reported at individual sites within some of these same locations. There 
have also been reports of more stable percent cover trends (e.g., 
Bonaire) and periods of increase (e.g., Flower Garden Banks). Observed 
declines in total coral cover in the Caribbean, since the major decline 
of Acropora spp. in the 1980s, have often been a result of the decline 
of the O. annularis species complex since the taxa can make up a large 
proportion of the total coral cover. Despite decreases, the O. 
annularis species complex continues to be reported as the dominant 
coral taxa, albeit at times its relative dominance has decreased to a 
lower percentage of the total coral cover (e.g., Curacao, U.S. Virgin 
Islands).
Other Biological Information
    The SRR and SIR provided the following information on Orbicella 
life history. Orbicella spp. have growth rates of approximately 1 cm 
per year, ranging from 0.06 to 1.2 cm per year. They grow more slowly 
in deeper water and in less clear water. Large colonies have lower 
total mortality rates than juvenile and small colonies.
    All three species of the O. annularis complex are hermaphroditic 
broadcast spawners, with spawning concentrated on six to eight nights 
following the full moon in late August, September, or early October. 
Orbicella faveolata is

[[Page 53937]]

largely reproductively incompatible with O. franksi and O. annularis, 
and it spawns about one to two hours earlier. Fertilization success 
measured in the field was generally below 15 percent for all three 
species being closely linked to the number of colonies concurrently 
spawning. In Puerto Rico, minimum size at reproduction for the O. 
annularis species complex was 83 cm\2\.
    Successful recruitment by the O. annularis species complex species 
has seemingly always been rare. Only a single recruit of Orbicella was 
observed over 18 years of intensive observation of 12 m\2\ of reef in 
Discovery Bay, Jamaica. Many other studies throughout the Caribbean 
also report negligible to absent recruitment of the species complex.
    The public comments did not provide new or supplemental information 
on the life history of Orbicella. Supplemental information we found on 
the life history of Orbicella includes the following. Orbicella franksi 
spawns an average of 110 minutes before O. annularis, and 120 minutes 
before O. faveolata (Fogarty et al., 2012a). Gametes can disperse over 
500 m in 100 minutes, and O. franksi sperm viability decreases after 
two hours (Levitan et al., 2004). Orbicella franksi and O. annularis 
gametes are compatible, though other mechanisms associated with the 
temporal isolation of spawning, including gamete aging, dilution, and 
dispersal, make hybridization less likely (Knowlton et al., 1997; 
Levitan et al., 2004). All three species are largely self-incompatible 
(Knowlton et al., 1997; Szmant et al., 1997). Size at sexual maturity 
is generally about 200 cm\2\ (Szmant-Froelich, 1985). Colonies that 
were fragmented experimentally to sizes smaller than 100 cm\2\ were 
generally found to have lower fecundity indicating that frequent 
fragmentation and partial mortality can affect reproductive capacity 
(Szmant-Froelich 1985).
    Smith and Aronson (2006) reported 18 Orbicella recruits between 
1998 and 2003 in 384 permanent monitoring quadrats (237 m\2\) in the 
lower Florida Keys. The ability of the species complex to dominate with 
such low recruitment rates has been described as a storage effect 
whereby large, old colonies are able to persist and maintain the 
population until favorable conditions for recruitment occur (Edmunds 
and Elahi, 2007). However, potential problems may exist for species 
employing storage effects if favorable conditions for recruitment occur 
so infrequently that they fall outside the life span of the cohort 
(Foster et al., 2013).
    All sources of information are used to summarize Orbicella's life 
history as follows. Orbicella species have slow growth rates, late 
reproductive maturity, and low recruitment rates. Colonies can grow 
very large and live for centuries. Large colonies have lower total 
mortality than small colonies, and partial mortality of large colonies 
can result in the production of ramets. The historical absence of small 
colonies and few observed recruits, even though large numbers of 
gametes are produced on an annual basis, suggests that recruitment 
events are rare and were less important for the survival of the O. 
annularis species complex in the past (Bruckner, 2012a). Large colonies 
in the species complex maintain the population until conditions 
favorable for recruitment occur; however, poor conditions can influence 
recruitment periodicity. While the life history strategy of the O. 
annularis species complex has allowed the taxa to remain abundant, we 
conclude that the buffering capacity of this life history strategy has 
been reduced by recent population declines and partial mortality, 
particularly in large colonies.
    The SRR, SIR, and public comments did not provide other biological 
information on the Orbicella annularis species complex. Supplemental 
biological information we found on Orbicella is provided as follows. 
The Orbicella annularis species complex is sensitive to cold water. In 
laboratory experiments, O. annularis species complex released 
zooxanthellae when shocked with cold water between 12 and 18 degrees C, 
and the response decreased with increasing temperature (Muscatine et 
al., 1991).
Susceptibility to Threats
    The SRR and SIR provided the following information on Orbicella's 
susceptibility to ocean warming. The Orbicella annularis species 
complex is moderately to highly susceptible to bleaching. The 
composition of zooxanthellae in at least some areas changes in response 
to bleaching. Bleaching has been shown to prevent reproduction in the 
following season after recovering normal pigmentation. Particularly 
well documented mortality following severe mass bleaching in 2005 
highlights the immense impact thermal stress events and their aftermath 
can have on the Orbicella annularis species complex. A significant 
correlation was found between bleaching in 2005 and the prevalence of 
yellow band disease and white plague affecting the Orbicella species 
complex. Additionally, in laboratory experiments, mortality due to 
yellow band disease increased with increasing temperatures.
    The public comments did not provide new or supplemental information 
on the susceptibility of Orbicella to ocean warming. Supplemental 
information we found on Orbicella's susceptibility to ocean warming 
confirms and expands the information in the SRR and SIR. The O. 
annularis species complex often has one of the highest bleaching levels 
among reported species. Extended recovery times have been reported, and 
disease outbreaks have often followed bleaching events. On Carysfort 
Reef in the Florida Keys, greater than 90 percent of O. annularis 
species complex colonies were bleached in March 1988 after the 1987 
Caribbean bleaching event; however, no colony mortality was observed 
between 1986 and 1988 (Fitt et al., 1993). Colonies of the O. annularis 
species complex in the Florida Keys that remained bleached seven months 
following the 1987 bleaching event experienced reproductive failure 
during the reproductive season following the bleaching event. Colonies 
that recovered after bleaching events were able to follow a normal 
reproductive cycle, but bleached colonies of O. annularis species 
complex were unable to complete gametogenesis (Szmant and Gassman, 
1990). Compared to recovered colonies, bleached colonies had lower 
tissue biomass, lower carbon-to-nitrogen ratios, and reduced growth, 
indicating the energy reserves needed for successful reproduction were 
not available (Szmant and Gassman, 1990).
    During the 1987 bleaching event, 90 percent of all O. annularis 
species complex colonies surveyed at 30 m in the Cayman Islands were 
bleached. Bleaching was less severe at 46 m with 14 percent of O. 
annularis species complex colonies bleached. Five months after 
bleaching was first observed in the Cayman Islands, 54 percent of 
bleached O. annularis species complex colonies had not recovered. 
Orbicella annularis species complex had the slowest recovery of the 28 
coral species observed to bleach (Ghiold and Smith, 1990).
    In a 1995 bleaching event in Belize, O. annularis species complex 
was the most affected coral taxon with 76 percent of the 2,126 surveyed 
colonies affected. Seven percent of the 904 colonies surveyed six 
months after the bleaching event remained bleached. Twenty-six percent 
of tagged O. annularis species complex colonies (n=19) exhibited 
partial mortality due to bleaching or post-bleaching infection by black 
band disease (McField, 1999).
    In 20 surveys across 302 sites throughout the wider Caribbean, O.

[[Page 53938]]

annularis species complex and Agaricia tenuifolia were the taxa most 
impacted by the 1998 bleaching event (Ginsburg and Lang, 2003; Kramer, 
2003). Subsequent disease outbreaks were also recorded in O. annularis 
and O. faveolata off Cura[ccedil]ao, the Cayman Islands, Costa Rica, 
and some of the Virgin Islands after the bleaching event. Bleaching and 
disease related mortality heavily impacted the O. annularis species 
complex (Ginsburg and Lang, 2003).
    During the 2005 bleaching event, approximately 70 percent of O. 
annularis species complex colonies bleached both in sites less than 10 
m in depth and in sites greater than 15 m in depth on the west and 
southwest coasts of Barbados (Oxenford et al., 2008). Bleaching was 
observed in 2005 at 86 of 94 sites (91 percent) surveyed in Buck Island 
Reef, U.S. Virgin Islands. Ninety-four percent of the cover of O. 
annularis species complex bleached (Clark et al., 2009).
    The 2005 bleaching event resulted in a 51 percent decrease in the 
cover of O. annularis species complex at five sites in the U.S. Virgin 
Islands between 2005 and 2007. Bleaching occurred in 16 of the 21 
species of coral at the five sites with maximum tissue area bleached 
between 98 to 99.5 percent for the O. annularis species complex. 
Mortality after the bleaching event occurred primarily from a 
subsequent regional outbreak of coral disease, predominantly white 
plague, not the bleaching itself. The highest rate of mortality of the 
19 species affected by the white plague was the Orbicella annularis 
species complex with 94.5 percent of disease lesions occurring on 
Orbicella annularis species complex. Total coral cover declined from 21 
percent to 10 percent, and species-specific changes in coral cover 
affected the relative abundance of coral species on the reef. Overall 
relative abundance of O. annularis species complex declined from an 
initial average of 79 to 59 percent of live coral cover (Miller et al., 
2009).
    Stratified random surveys on back-reefs and fore-reefs between one 
and 30 m depth off Puerto Rico (Mona and Desecho Islands, La Parguera, 
Mayaguez, Boqueron, and Rincon) in 2005 and 2006 revealed bleaching was 
most severe in O. annularis species complex with 94 percent of colonies 
bleached. After bleaching, a disease outbreak occurred, and O. 
annularis species complex suffered extensive partial and total 
mortality. Coral cover declined between 40 and 60 percent and was 
primarily driven by mortality of O. annularis species complex. 
Additionally, the severe tissue loss and prolonged bleaching stress 
resulted in reproductive collapse of O. annularis species complex 
during the 2006 mass spawning cycle (Waddell and Clarke, 2008).
    The 2005 bleaching affected greater than 95 percent of O. annularis 
species complex in Mona and Desecheo Islands, Puerto Rico and was 
followed by a disease outbreak that both caused extensive mortality 
(Bruckner and Hill, 2009). A study of 36 sites across six countries 
(Grenada, Cura[ccedil]ao, Panam[aacute], Puerto Rico, Cayman Islands, 
and Bermuda) and three depth habitats (less than 4 m, 5 to 12 m, and 
greater than 15 m) found a significant correlation between the 2005 
bleaching and prevalence of yellow band disease and white plague in O. 
annularis species complex (Croquer and Weil, 2009). Orbicella annularis 
species complex bleached at all depths surveyed in Grenada (23 to 52 
percent of colonies), Puerto Rico (21 to 40 percent), and Cayman 
Islands (16 to 44 percent). The species complex did not experience 
bleaching in Curacao or Bermuda, both locations reported very low 
bleaching across all genera examined (Croquer and Weil, 2009). 
Bleaching of O. annularis species complex varied by depth in Panama 
with bleaching occurring in 11 percent of colonies in depths less than 
4 m and in15 percent of colonies in depths between 5 and 12 m, but no 
bleaching occurred in deep depths greater than 15 m (Croquer and Weil, 
2009). Smith et al. (2013b) described species responses to the 2005 and 
2010 bleaching events in St. Thomas, St. Croix, and St. John, U.S. 
Virgin Islands. The response of the O. annularis species complex 
(mostly O. faveolata and O. franksi with the likelihood of small 
numbers of O. annularis) to the 2005 bleaching event was high to 
moderate initial response of bleaching prevalence, high disease 
prevalence, high mortality, a large decline in coral cover, and 
increasing or stable colony abundance. Average bleaching was 66 
percent, and paling was 27 percent in 2005. Disease prevalence in O. 
annularis complex was 17 percent after the 2005 bleaching event. In the 
milder 2010 bleaching event, 35 percent of O. annularis species complex 
colonies bleached, and 47 percent of O. annularis species complex 
colonies paled. Less than one percent of O. annularis species complex 
colonies suffered total mortality, but percent cover decreased from 
seven percent cover of O. annularis species complex in 2005 before 
bleaching to less than three percent in 2007. By 2010, there was a 
slight increase in percent cover to about four percent. Orbicella 
annularis species complex lost a large proportion of colonies in the 
largest size class and showed a significant increase in colony 
abundance, likely due to the increase in abundance of colonies in 
smaller size classes resulting from partial mortality of larger 
colonies.
    Van Woesik et al. (2012) developed a coral resiliency index based 
on biological traits and processes to evaluate extinction risk due to 
bleaching. Evaluations were performed at the genus level, but genera 
were separated between the Caribbean and Indo-Pacific. They rated the 
resilience score for the O. annularis species complex as four out of a 
range of -6 to 7 observed in other coral genera. Less than or equal to 
-3 was considered highly vulnerable to extinction, and greater than or 
equal to 4 was considered highly tolerant. Thus, O. annularis species 
complex was rated as highly tolerant. However, Smith et al. (2013b) 
concluded that large faviids, such as the O. annularis species complex, 
seem very susceptible to long-term population declines because of their 
poor response to stress response when bleaching, disease, and mortality 
were considered. The O. annularis species complex was found to be 
likely less equipped to recovery after bleaching because they tend to 
grow slowly, have lower fecundity, and are more susceptible to 
mortality when small (Smith et al., 2013b). While the van Woesik et al. 
(2012) study was in the SIR, the findings specific to Orbicella were 
not included. The public comments indicated the results of this study 
should be considered in the listing status of the three species in the 
Orbicella species complex.
    All sources of information are used to describe Orbicella's 
susceptibility to ocean warming as follows. The O. annularis species 
complex is highly susceptible to ocean warming. Bleaching often occurs 
in 76 to 94 percent of O. annularis species complex colonies during 
bleaching events, and Orbicella spp. are one of the taxa most affected 
by high temperatures. Colonies in deeper water have been reported to 
bleach less severely. Recovery from bleaching can take longer for the 
species complex than for other coral species, and prolonged stress from 
bleaching has been cited as a possible reason for reproductive failure 
following bleaching events. Mortality from temperature anomalies is 
often due to subsequent disease outbreaks. Thus, we conclude that the 
O. annularis species complex is highly susceptible to ocean warming.
    The SRR and SIR provided the following information on Orbicella's

[[Page 53939]]

susceptibility to acidification. The only study conducted regarding the 
impact of acidification on this genus is a field study that did not 
find any change in O. faveolata calcification in field-sampled colonies 
from the Florida Keys up through 1996. Preliminary experiments testing 
effects of acidification on fertilization and settlement success of O. 
annularis species complex show results that are consistent with the 
significant impairments demonstrated for A. palmata.
    The public comments did not provide new or supplemental information 
on the susceptibility of the Orbicella species complex to 
acidification. Supplemental information we found on the susceptibility 
of the Orbicella species complex to acidification includes the 
following. In laboratory experiments, reproduction of O. faveolata was 
negatively impacted by increasing carbon dioxide, and impairment of 
fertilization was exacerbated at lower sperm concentrations (Albright, 
2011b). Fertilization success was reduced by 25 percent at 529 [mu]atm 
(43 percent fertilization) and 40 percent at 712 [mu]atm (34 percent 
fertilization) compared to controls at 435 [mu]atm (57 percent 
fertilization; Albright, 2011a). Additionally, growth rate of O. 
faveolata was reduced under lower pH conditions (7.6) compared to 
higher pH conditions (8.1) after 120 days of exposure (Hall et al., 
2012).
    All sources of information are used to describe Orbicella's 
susceptibility to acidification as follows. Laboratory studies indicate 
that Orbicella is susceptible to ocean acidification both through 
reduced fertilization of gametes and reduced growth of colonies. Thus, 
we conclude that the Orbicella species complex is highly susceptible to 
ocean acidification.
    The SRR and SIR provided the following information on Orbicella's 
susceptibility to disease. White plague and yellow band (also called 
yellow blotch) disease have caused profound population decline of the 
Orbicella annularis species complex both with and without prior 
bleaching.
    The public comments did not provide new or supplemental information 
on the susceptibility of Orbicella spp. to disease. Supplemental 
information we found on Orbicella's susceptibility to disease confirms 
and expands the information in the SRR and SIR. Orbicella spp. are 
susceptible to black band disease and dark spot syndrome (Alcolado et 
al., 2010). Additionally, an unknown disease was observed in the Red 
Hind Marine Conservation District in the U.S. Virgin Islands and 
affected 39 percent of O. annularis species complex colonies (Smith et 
al., 2010). White plague is one of the most aggressive coral diseases 
in the Caribbean with progression rates of 1 to 10 cm per day (Bruckner 
and Hill, 2009). Tissue loss from yellow band disease is slow, 
averaging 0.5 to 1 cm per month, though tissue loss can be significant 
over the long term since colonies can remain infected for years and can 
have multiple lesions per colony (Bruckner and Bruckner, 2006b).
    In the Florida Keys, the prevalence of white plague increased 
between 1996 and 2002. No O. annularis species complex colonies with 
white plague were reported within monitoring stations in 1996, but 
infected colonies appeared in 32 stations in 2002 (Waddell, 2005). 
Orbicella annularis species complex had the highest prevalence (up to 
12 percent) of the 21 species affected by white plague in Puerto Rico 
between 1998 and 2008 (Bruckner and Hill, 2009). In Mexico, O. 
annularis species complex had the highest disease prevalence in surveys 
during 2004 (27 percent, Ward et al., 2006). Surveys in four locations 
(Netherlands Antilles, Grenada, Turks and Caicos, and U.S. Virgin 
Islands) between 1997 and 1998 revealed that prevalence of yellow band 
in O. annularis species complex ranged from 18 to 91 percent.
    Tagged colonies with yellow band disease in Puerto Rico lost an 
average of 32 percent of their tissue over four years, and the percent 
of partial mortality appeared to increase with colony size (Bruckner 
and Bruckner, 2006b). Eight percent of infected colonies died 
completely (most were 50 cm or less in size), and larger colonies lost 
between 60 and 85 percent of their tissue (Bruckner and Bruckner, 
2006b). Eighty-five percent of colonies with yellow band disease tagged 
in 1999 still had active signs of the disease in 2003 (Bruckner and 
Bruckner, 2006b). In 1999, yellow band disease affected up to 50 
percent of all O. annularis species complex colonies at permanent sites 
in Puerto Rico, including many of the largest (2 to 3 m diameter and 
height) and presumably oldest colonies (Waddell and Clarke, 2008).
    In Curacao, yellow band disease affected from three to 49 percent 
of all O. annularis species complex colonies within transects conducted 
on western reefs between 1997 and 2005. The highest prevalence of 
yellow band disease occurred in 1997 and 1998. Thirty-one to 49 percent 
of O. annularis species complex colonies were affected in eastern 
Curacao, and 24 percent were affected in western Curacao. The numbers 
of new infections declined from 2000 to 2005. Yellow band disease 
affected larger corals more frequently than small corals. Over 21 
percent of the colonies tagged with yellow band disease between 1997 
and 1999 were still infected in 2005. Of the remainder, 44 percent 
died, 2 percent were affected by other diseases, and 32 percent no 
longer had signs of yellow band disease but had large amounts (most 
greater than 90 percent) of partial mortality (Bruckner and Bruckner, 
2006a).
    Disease prevalence in O. annularis species complex (O. annularis 
and O. faveolata) at three reefs off Mexico increased from between zero 
and four percent in 1996 and 1998 to between 26 and 37 percent in 2001. 
The increase was due to the proliferation of yellow band disease, 
though black band disease and white plague were also present. Partial 
mortality also increased over this same period from 20 to 35 percent of 
O. annularis species complex colonies at one site and from 35 to 52 
percent at another (Jordan-Dahlgren et al., 2005).
    At 253 sites surveyed in 2009 in St. Croix and St. John, U.S. 
Virgin Islands and La Parguera, Puerto Rico, the average number of 
healthy O. annularis species complex colonies was 182 ( 33 
SE) per 100 m\2\. Yellow band was present on an average of about one 
percent of colonies (Muller and van Woesik, 2012).
    All sources of information are used to describe Orbicella's 
susceptibility to disease as follows. Disease can affect a large 
proportion of the Orbicella spp. population (3 to 91 percent), 
particularly during outbreaks following bleaching events, and can cause 
extensive mortality. Partial mortality can be high (32 to greater than 
90 percent) and can result in multiple ramets. White plague and yellow 
band disease have had the greatest effect and can disproportionately 
affect larger colonies in the species complex. Total colony mortality 
is less likely for larger colonies than for smaller colonies, and 
partial mortality can lead to changes in colony size distribution as 
observed in Puerto Rico, U.S. Virgin Islands, and a study in Bahamas, 
Bonaire, Cayman Islands, and St. Kitts and Nevis. Thus, we conclude 
that the O. annularis species complex is highly susceptible to disease.
    The SIR and SRR did not provide any information on the trophic 
effects of fishing on Orbicella. The public comments did not provide 
new or supplemental information, and we did not find new or 
supplemental information on the trophic effects of fishing on 
Orbicella. However, as described above in Caribbean Genera and 
Species--Introduction, due to the

[[Page 53940]]

level of reef fishing conducted in the Caribbean, coupled with Diadema 
die-off and lack of significant recovery, competition with algae can 
adversely affect coral recruitment. This effect coupled with 
Orbicella's low recruitment rate indicates it likely has some 
susceptibility to the trophic effects of fishing. The available 
information does not support a more precise description of its 
susceptibility.
    The SRR and SIR provided the following information on the 
susceptibility of Orbicella to sedimentation. Orbicella has shown a 
decline in growth at sediment impacted sites in Puerto Rico and during 
periods of construction in Aruba. Along a gradient of continental 
influence in the southern Gulf of Mexico, density and calcification 
rate of O. annularis decreased with increasing turbidity and 
sedimentation while extension rate increased with increasing turbidity 
and sedimentation.
    The public comments did not provide new or supplemental information 
on the susceptibility of Orbicella to sedimentation. Supplemental 
information we found on the susceptibility of the Orbicella annularis 
species complex confirms the information in the SRR and SIR. The 
Orbicella annularis species complex appears to be moderately capable of 
removing sediment from the colony surface. Colonies receiving single 
applications of 200 or 400 mg sediment per cm\2\ showed no evidence of 
damage while 800 mg per cm\2\ caused mortality (Rogers, 1983). 
Sedimentation has been found to negatively affect O. annularis species 
complex primary production, growth rates, and abundance (Pastorok and 
Bilyard, 1985). An observed difference in average colony size at two 
sites in Puerto Rico led Loya (1976) to conclude turbidity negatively 
affects growth of O. annularis species complex since colony size was 
half as large at the sediment-impacted site (23 cm versus 9 cm).
    All sources of information are used to describe Orbicella's 
susceptibility to sedimentation as follows. Although the species 
complex is moderately capable of removing sediment from the colony 
surface, sedimentation negatively affects primary production, growth 
rates, calcification, colony size, and abundance. Thus, we conclude 
that the O. annularis species complex is highly susceptible to 
sedimentation.
    The SRR and SIR provided the following information on the 
susceptibility of Orbicella to nutrient enrichment. Orbicella had an 
increasing growth rate with improving environmental conditions in 
Barbados. Additionally, decreasing growth rate of Orbicella over a 30-
year period was attributed to deterioration of water quality.
    The public comments did not provide new or supplemental information 
on the susceptibility of Orbicella to nutrient enrichment. Supplemental 
information we found on the susceptibility of the Orbicella species 
complex confirms and expands the information in the SRR and SIR. Two 
growth forms of O. annularis species complex, columnar (likely O. 
faveolata) and lobate (likely O. annularis) were found to have 
increasing average growth rates with improving environmental conditions 
away from a eutrophication gradient in Barbados (Tomascik, 1990). 
Although nutrient concentration was negatively correlated with growth, 
suspended particulate matter resulting from eutrophication, rather than 
the nutrients themselves, was postulated to be the cause of observed 
decreased growth rates (Tomascik and Sander, 1985). A general pattern 
of decreasing growth rates of the columnar growth form between 1950 and 
1983 may be directly related to the deterioration of water quality 
along the west coast of the island (Tomascik, 1990). Additionally, 
Orbicella spp. did not recruit to settlement plates on the most 
eutrophic reef, and recruitment of Orbicella spp. increased at sites 
with decreasing eutrophication along the eutrophication gradient 
(Tomascik, 1991). Field experiments indicate that nutrient enrichment 
significantly increases yellow band disease severity in O. annularis 
and O. franksi through increased tissue loss (Bruno et al., 2003).
    All sources of information are used to describe Orbicella's 
susceptibility to nutrient enrichment as follows. The Orbicella 
annularis species complex is susceptible to nutrient enrichment through 
reduced growth rates, lowered recruitment, and increased disease 
severity. Thus, we conclude that the O. annularis species complex is 
highly susceptible to nutrient enrichment.
    The SRR and SIR provided the following information on the 
susceptibility of Orbicella to predation. Predators of the O. annularis 
species complex include the corallivorous snail Coralliophila 
abbreviata and some species of parrotfish including Sparisoma viride 
and S. aurofrenatum. Additionally, damselfish remove live coral tissue 
to build algal gardens. The large decline of Acropora spp. in the 
Caribbean, likely resulted in greater impacts by damselfishes on other 
high-dimension corals, including the O. annularis species complex.
    Public comments did not provide new or supplemental information on 
the susceptibility of Orbicella to predation. Supplemental information 
we found on the susceptibility of the Orbicella species complex 
includes the following. Surveys of six sites in Navassa found between 
zero and 33 percent of O. annularis species complex colonies (average 
17 percent across all sites) were affected by C. abbreviata (Miller et 
al., 2005). The O. annularis species complex was the preferred target 
of parrotfish across all reef habitats in a study on the Belize barrier 
reef. Incidence of parrotfish grazing was highest on O. annularis (over 
55 percent of colonies), followed by O. franksi and O. faveolata, 
respectively (Rotjan, 2007). In most habitats, a few colonies of 
Orbicella spp. were more heavily grazed by parrotfishes, while the 
majority showed little or no parrotfish grazing (Rotjan and Lewis, 
2006).
    All sources of information are used to describe Orbicella's 
susceptibility to predation as follows. The O. annularis species 
complex is susceptible to several predators. Current effects of 
predation appear to be low. Thus, we conclude the O. annularis species 
complex has low susceptibility to predation.
    The SRR and SIR did not provide information on the effects of sea 
level rise on Orbicella. The SRR described sea level rise as an overall 
low to medium threat for all coral species. The public comments did not 
provide new or supplemental information on Orbicella's susceptibility 
to sea level rise, and we did not find any new or supplemental 
information. Thus, we conclude that Orbicella has some susceptibility 
to sea level rise, but the available information does not support a 
more precise description of susceptibility to this threat.
    The SRR and SIR provided the following information on the 
susceptibility of the Orbicella species complex to collection and 
trade. The Orbicella complex species have a very low occurrence in the 
CITES trade databases. Hence, collection and trade is not considered a 
significant threat to the Orbicella annularis complex species. The 
public comments did not provide new or supplemental information on the 
susceptibility of the Orbicella species complex to trade. Supplemental 
information we found on the susceptibility of species in the Orbicella 
complex to collection and trade is described in each of the individual 
species sections.

[[Page 53941]]

Genus Conclusion
    The O. annularis species complex is distributed throughout the 
Caribbean and occupies a variety of habitats across a large depth 
range, including mesophotic depths to 90 m. Over the last twenty years, 
major declines of approximately 50 to 95 percent have occurred. In 
addition, changes in size frequency distribution have sometimes 
accompanied decreases in cover, resulting in fewer large colonies that 
impact the buffering capacity of the species complex's life history 
strategy. Despite decline, the O. annularis species complex continues 
to be reported as the dominant coral taxon, sometimes at a lower 
percentage of the total coral cover.
    The species complex has highly susceptibility to ocean warming, 
acidification, disease, sedimentation, and nutrients; some 
susceptibility to trophic effects of fishing and sea level rise; and 
low susceptibility to predation. Susceptibility to collection and trade 
is described in each of the individual species sections.

Orbicella faveolata

Introduction
    The SRR and SIR provided the following information on O. 
faveolata's morphology. Orbicella faveolata grows in heads or sheets, 
the surface of which may be smooth or have keels or bumps. The skeleton 
is much less dense than in the other two Orbicella species. Colony 
diameter can reach up to 10 m with a height of 4 to 5 m. The public 
comments did not provide new or supplemental information on O. 
faveolata's morphology, and we did not find any new or supplemental 
information.
Spatial Information
    The SRR and SIR provided the following information on the 
distribution, habitat and depth range of O. faveolata. Orbicella 
faveolata occurs in the western Atlantic and throughout the Caribbean, 
including Bahamas, Flower Garden Banks, and the entire Caribbean 
coastline. There is conflicting information on whether or not it occurs 
in Bermuda. Orbicella faveolata has been reported in most reef habitats 
and is often the most abundant coral between 10 and 20 m in fore-reef 
environments. The depth range of O. faveolata has been reported as 0.5 
to 40 m, though the species complex has been reported to depths of 90 
m, indicating O. faveolata's depth distribution is likely deeper than 
40 m. Orbicella species are a common, often dominant component of 
Caribbean mesophotic reefs, suggesting the potential for deep refugia 
for O. faveolata.
    The public comments did not provide new or supplemental information 
on O. faveolata's distribution, habitat, or depth range. Supplemental 
information we found includes the following. Veron (2014) confirmed the 
occurrence of O. faveolata in five out of his 11 ecoregions in the west 
Atlantic and greater Caribbean known to contain corals and strongly 
predicted its presence in an additional three ecoregions (off Colombia 
and Venezuela; Jamaica and Cayman Islands; and Florida and the 
Bahamas). Many studies have confirmed the presence of O. faveolata in 
these additional three ecoregions (Bayraktarov et al., 2012; Bruckner, 
2012a; Burman et al., 2012). The ecoregions where Veron (2014) reported 
the absence of O. faveolata are off the coasts of Brazil, Bermuda, and 
the southeastern U.S. north of southern Florida (Veron, 2014). Smith 
(2013) reported that O. faveolata is found in the U.S. Virgin Islands 
across all depths to about 45 m.
Demographic Information
    The SRR and SIR provided the following information on O. 
faveolata's abundance and population trends. Orbicella faveolata is 
considered common.
    The public comments did not provide new or supplemental information 
on O. faveolata's population trends but provided the following 
supplemental information on O. faveolata's abundance. Extrapolated 
population estimates from stratified random samples in the Florida Keys 
were 39.7  8 million (SE) colonies in 2005, 21.9  7 million (SE) colonies in 2009, and 47.3  14.5 
million (SE) colonies in 2012. The greatest proportion of colonies 
tended to fall in the 10 to 20 cm and 20 to 30 cm size classes in all 
survey years, but there was a fairly large proportion of colonies in 
the greater than 90 cm size class. Partial mortality of the colonies 
was between 10 and 60 percent surface across all size classes. In the 
Dry Tortugas, Florida, O. faveolata ranked seventh most abundant out of 
43 coral species in 2006 and fifth most abundant out of 40 in 2008. 
Extrapolated population estimates were 36.1  4.8 million 
(SE) colonies in 2006 and 30  3.3 million (SE) colonies in 
2008. The size classes with the largest proportion of colonies were 10 
to 20 cm and 20 to 30 cm, but there was a fairly large proportion of 
colonies in the greater than 90 cm size class. Partial mortality of the 
colonies ranged between approximately two percent and 50 percent. 
Because these population abundance estimates are based on random 
surveys, differences between years may be attributed to sampling effort 
rather than population trends (Miller et al., 2013).
    Supplemental information we found on O. faveolata's abundance and 
population trends includes the following. In a survey of 31 sites in 
Dominica between 1999 and 2002, O. faveolata was present at 80 percent 
of the sites at one to ten percent cover (Steiner, 2003). In a 1995 
survey of 16 reefs in the Florida Keys, O. faveolata ranked as the 
coral species with the second highest percent cover (Murdoch and 
Aronson, 1999). On 84 patch reefs (3 to 5 m depth) spanning 240 km in 
the Florida Keys, O. faveolata was the third most abundant coral 
species comprising seven percent of the 17,568 colonies encountered and 
was present at 95 percent of surveyed reefs between 2001 and 2003 
(Lirman and Fong, 2007). In surveys of 280 sites in the upper Florida 
Keys in 2011, O. faveolata was present at 87 percent of sites visited 
(Miller et al., 2011b). In 2003 on the East Flower Garden Bank, O. 
faveolata comprised ten percent of the 76.5 percent coral cover on 
reefs 32 to 40 m, and partial mortality due to bleaching, disease, and 
predation were rare at monitoring stations (Precht et al., 2005).
    Colony density ranges from approximately 0.1 to 1.8 colonies per 10 
m\2\ and varies by habitat and location. In surveys along the Florida 
reef tract from Martin County to the lower Florida Keys, density of O. 
faveolata was approximately 1.6 colonies per 10 m\2\ (Wagner et al., 
2010). On remote reefs off southwest Cuba, density of O. faveolata was 
0.12  0.20 (SD) colonies per 10 m transect on 38 reef-crest 
sites and 1.26  1.06 colonies per 10 m transect on 30 reef-
front sites (Alcolado et al. 2010). In surveys of 1,176 sites in 
southeast Florida, the Dry Tortugas, and the Florida Keys between 2005 
and 2010, density of O. faveolata ranged between 0.17 and 1.75 colonies 
per 10 m\2\ and was highest on mid-channel reefs followed by offshore 
patch reefs and fore-reefs (Burman et al., 2012). Along the east coast 
of Florida, density was highest in areas south of Miami at 0.94 
colonies per 10 m\2\ compared to 0.11 colonies per 10 m\2\ in Palm 
Beach and Broward Counties (Burman et al., 2012).
    Orbicella faveolata is the sixth most abundant species by percent 
cover in permanent monitoring stations in the U.S. Virgin Islands. The 
species complex had the highest abundance and included all colonies 
where species identification was uncertain. Therefore, O. faveolata is 
likely more abundant. Population estimates in the 49 km\2\ Red

[[Page 53942]]

Hind Marine Conservation District are at least 16 million colonies 
(Smith, 2013).
    Population trend data exists for several locations. At nine sites 
off Mona and Desecheo Islands, Puerto Rico, no species extirpations 
were noted at any site over ten years of monitoring between 1998 and 
2008 (Bruckner and Hill, 2009). Both O. faveolata and O. annularis 
sustained the large losses during the period. The number of colonies of 
O. faveolata decreased by 36 and 48 percent at Mona and Desecheo 
Islands, respectively (Bruckner and Hill, 2009). In 1998, 27 percent of 
all corals at six sites surveyed off Mona Island were O. faveolata 
colonies, but decreased to approximately 11 percent in 2008 (Bruckner 
and Hill, 2009). At Desecheo Island, 12 percent of all coral colonies 
were O. faveolata in 2000 compared to seven percent in 2008.
    In a survey of 185 sites in five countries (Bahamas, Bonaire, 
Cayman Islands, Puerto Rico, and St. Kitts and Nevis) between 2010 and 
2011, size of O. faveolata colonies was significantly greater than O. 
franksi and O. annularis. The total mean partial mortality of O. 
faveolata at all sites was 38 percent. The total live area occupied by 
O. faveolata declined by a mean of 65 percent, and mean colony size 
declined from 4005 cm\2\ to 1413 cm\2\. At the same time, there was a 
168 percent increase in small tissue remnants less than 500 cm\2\, 
while the proportion of completely live large (1,500 to 30,000 cm\2\) 
colonies decreased. Orbicella faveolata colonies in Puerto Rico were 
much larger and sustained higher levels of mortality compared to the 
other four countries. Colonies in Bonaire were also large but 
experienced much lower levels of mortality. Mortality was attributed 
primarily to outbreaks of white plague and yellow band disease, which 
emerged as corals began recovering from mass bleaching events. This was 
followed by increased predation and removal of live tissue by 
damselfish to cultivate algal lawns (Bruckner, 2012a).
    All information on O. faveolata's abundance and population trends 
can be summarized as follows. Orbicella faveolata is a common species 
throughout the greater Caribbean. Based on population estimates, there 
are at least tens of millions of colonies present in each of several 
locations including the Florida Keys, Dry Tortugas, and the U.S. Virgin 
Islands. Absolute abundance is higher than the estimate from these 
three locations given the presence of this species in many other 
locations throughout its range. Population decline has occurred over 
the past few decades with a 65 percent loss in O. faveolata cover 
across five countries. Losses of O. faveolata from Mona and Descheo 
Islands, Puerto Rico include a 36 to 48 percent reduction in abundance 
and a decrease of 42 to 59 percent in its relative abundance (i.e., 
proportion relative to all coral colonies). High partial mortality of 
colonies has led to smaller colony sizes and a decrease of larger 
colonies in some locations such as the Bahamas, Bonaire, Puerto Rico, 
Cayman Islands, and St. Kitts and Nevis. Partial colony mortality is 
lower in some areas such as the Flower Garden Banks. We conclude that 
O. faveolata has declined but remains common and likely has at least 
tens of millions of colonies throughout its range. Additionally as 
discussed in the genus section, we conclude that the buffering capacity 
of O. faveolata's life history strategy that has allowed it to remain 
abundant has been reduced by the recent population declines and amounts 
of partial mortality, particularly in large colonies.
Other Biological Information
    The SRR and SIR provided the following information on O. 
faveolata's life history. In many life history characteristics, 
including growth rates, tissue regeneration, and egg size, O. faveolata 
is considered intermediate between O. annularis and O. franksi. Spatial 
distribution may affect fecundity on the reef, with deeper colonies of 
O. faveolata being less fecund due to greater polyp spacing.
    The public comments did not provide new or supplemental information 
on the life history of O. faveolata. Supplemental information we found 
on O. faveolata's life history includes the following. Reported growth 
rates of O. faveolata range between 0.3 and 1.6 cm per year (Cruz-
Pi[ntilde][oacute]n et al., 2003; Tomascik, 1990; Villinski, 2003; 
Waddell, 2005). Graham and van Woesik (2013) report that 44 percent of 
small colonies of O. faveolata in Puerto Morelos, Mexico, resulting 
from partial colony mortality produced eggs at sizes smaller than 
maturation. The number of eggs produced per unit area of smaller 
fragments was significantly less than in larger size classes. Szmant 
and Miller (2005) reported low post-settlement survivorship for O. 
faveolata transplanted to the field with only three to 15 percent 
remaining alive after 30 days. Post-settlement survivorship was much 
lower than the 29 percent observed for A. palmata after seven months 
(Szmant and Miller, 2005). Darling et al. (2012) performed a biological 
trait-based analysis to categorize coral species into four life history 
strategies: Generalist, weedy, competitive, and stress-tolerant. The 
classifications were primarily separated by colony morphology, growth 
rate, and reproductive mode. Orbicella faveolata was classified as a 
``generalist'' species, thus likely less vulnerable to environmental 
stress.
    The SRR and SIR provided the following other biological information 
on O. faveolata. Surveys at an inshore patch reef in the Florida Keys 
that experienced temperatures less than 18 degrees C for 11 days 
revealed species-specific cold-water susceptibility and survivorship. 
Orbicella faveolata was one of the more susceptible species with 90 
percent of colonies experiencing total colony mortality, including some 
colonies estimated to be more than 200 years old (Kemp et al., 2011). 
In surveys from Martin County to the lower Florida Keys, O. faveolata 
was the second most susceptible coral species experiencing an average 
of 37 percent partial mortality (Lirman et al., 2011).
    The public comments did not provide any new or supplemental 
biological information on O. faveolata. Supplemental biological 
information we found on O. faveolata includes the following. Samples (n 
= 182) of O. faveolata from the upper and lower Florida Keys and Mexico 
showed three well-defined populations based on five genetic markers, 
but the populations were not stratified by geography, indicating they 
were shared among the three regions (Baums et al., 2010). Of ten O. 
faveolata colonies observed to spawn at a site off Bocas del Toro, 
Panama, colonies sorted into three spatially arranged genotypes 
(Levitan et al., 2011).
    Orbicella faveolata larvae are sensitive to ultraviolet radiation 
during the motile planula stage through the onset of larval competence 
(Aranda et al., 2011). Of six Caribbean coral species exposed to high 
solar irradiation, O. faveolata and Stephanocoenia intersepta had the 
most severe decline in photochemical efficiency resulting in severe 
tissue loss and mortality (Fournie et al., 2012).
    Experiments exposing O. faveolata to high temperatures (up to 35 
degrees C) revealed that the corals produced heat shock proteins at 
temperatures between 33 and 35 degrees C even for very short exposures 
(2 h) but did respond at temperatures between 27 and 31 degrees C when 
exposed from 2 hours to one week (Black et al., 1995).
    Thornhill et al. (2006) repeatedly sampled symbiont composition of 
colonies of six coral species in the Bahamas and the Florida Keys in 
1998 and 2000 to 2004, during and after the 1997-98 bleaching event. 
Symbioses in O. faveolata remained stable at virtually

[[Page 53943]]

all sites in the Bahamas and the Florida Keys. Individual colonies 
usually showed fidelity over time to one particular Symbiodinium 
partner, and changing symbiont types was rare, thus indicating 
acclimation to warming temperatures may not occur by symbiont 
shuffling.
Susceptibility to Threats
    The threat susceptibility information from the SRR and SIR was 
interpreted in the proposed rule for O. faveolata's vulnerabilities to 
threats as follows: High vulnerability to ocean warming, disease, 
acidification, sedimentation, and nutrient enrichment; moderate 
vulnerability to the trophic effects of fishing; and low vulnerability 
to sea level rise, predation, and collection and trade.
    The SRR and SIR provided the following information on the 
susceptibility of O. faveolata to ocean warming. Recent work in the 
Mesoamerican reef system indicated that O. faveolata had reduced 
thermal tolerance in locations with increasing human populations and 
over time, implying increasing local threats. At sites in Navassa, O. 
faveolata and Agaricia spp. were the most susceptible to bleaching. 
Approximately 90 percent of O. faveolata colonies (n = 334) bleached at 
deeper sites (>18 m), and approximately 60 percent of O. faveolata 
colonies (n = 20) bleached at shallower sites (<10 m) in 2006. During a 
moderate bleaching event in Colombia in 2010, 100 percent of O. 
faveolata colonies bleached at a site in Gayraca Bay, and 50 percent of 
O. faveolata colonies were dead and completely overgrown by algae in 
2011 (Bayraktarov et al., 2012).
    The public comments did not provide new or supplemental information 
on the susceptibility of O. faveolata to ocean warming. Supplemental 
information we found on the susceptibility of O. faveolata to ocean 
warming includes the following. Stratified random surveys on back-reefs 
and fore-reefs between one and 30 m depth off Puerto Rico (Mona and 
Desecho Islands, La Parguera, Mayaguez, Boqueron, and Rincon) in 2005 
and 2006 revealed severe bleaching in O. faveolata with approximately 
90 percent of colonies bleached (Waddell and Clarke, 2008). Surveys 
from 2005 to 2007 along the Florida reef tract from Martin County to 
the lower Florida Keys indicated that O. faveolata had the 13th highest 
bleaching prevalence out of 30 species observed to bleach (Wagner et 
al., 2010). During a 2009 bleaching event on Little Cayman, of the ten 
coral species that bleached, O. faveolata had the third highest 
bleaching prevalence with approximately 37 percent of colonies bleached 
(van Hooidonk et al., 2012).
    Coral cores from 92 colonies of O. faveolata from the Mesoamerican 
Reef around Belize and Honduras indicate that the bleaching event in 
1998 was unprecedented in the prior century despite periods of higher 
temperatures and solar irradiance (Carilli et al., 2010). The authors 
of the study concluded that bleaching in 1998 likely stemmed from 
reduced thermal tolerance due to the synergistic impacts of chronic 
local stressors stemming from land-based sources of pollution (Carilli 
et al., 2010). Coral cores collected from four sites in Belize indicate 
that O. faveolata that experienced higher chronic stress were more 
severely affected by bleaching and had a much slower recovery after the 
severe 1998 bleaching event (Carilli et al., 2009). Coral growth rates 
at sites with higher local anthropogenic stressors remained suppressed 
for at least eight years, while coral growth rates at sites with lower 
stress recovered in two to three years (Carilli et al., 2009). Based on 
samples of O. faveolata and O. franksi collected from the Mesoamerican 
Barrier Reef, calcification of these two species is projected to cease 
at 35 degrees C in this location, even without an increase in 
acidification (Carricart-Ganivet et al., 2012). Collections from 
Chinchorro Bank indicate that calcification of O. faveolata decreased 
20 percent over the period of 1985 to 2009 where there was a 0.6 degree 
C increase in sea surface temperature (equivalent to 2.4 degrees C per 
century; Carricart-Ganivet et al., 2012).
    Polato et al. (2010) raised O. faveolata larvae derived from three 
to four colonies from Florida and Mexico under mean and elevated (1 to 
2 degrees above summer mean) temperatures. Both locations had misshapen 
embryos at the elevated temperature, but the percentage was higher in 
the embryos from Florida. They found conserved and location-specific 
variation in gene expression in processes related to apoptosis 
(programmed cell death), cell structuring, adhesion and development, 
energy and protein metabolism, and response to stress.
    Voolstra et al. (2009) exposed O. faveolata embryos to temperatures 
of 27.5, 29, and 31.5 degrees C directly after fertilization and 
measured differences in gene expression after 12 and 48 hours. They 
found a higher number of misshapen embryos after 12 hours at 29 and 
31.5 degrees C in comparison to embryos kept at 27.5 degrees C. 
However, after 48 hours, the proportion of misshapen embryos decreased 
for embryos kept at 29 and 31.5 degrees C, and increased for embryos 
kept at 27.5 degrees C. Increased temperatures may lead to oxidative 
stress, apoptosis, and a structural reconfiguration of the cytoskeletal 
network. However, embryos responded differently depending on exposure 
time and temperature level. Embryos showed expression of stress-related 
genes at a temperature of 29 degrees C but seemed to be able to 
counteract the initial response over time. Embryos at 31.5 degrees C 
displayed continuous expression of stress genes.
    During the 2005 bleaching event, larger colonies of O. faveolata 
experienced more intensive bleaching than smaller colonies at inshore 
patch reefs of the Florida Keys (Brandt, 2009). Orbicella faveolata was 
one of the most affected species with approximately 80 percent of 
colonies (n = 77) bleached and, out of eight species that bleached, had 
the fourth highest bleaching prevalence (Brandt, 2009). Orbicella 
faveolata colonies with greater bleaching intensities later developed 
white plague disease (Brandt and McManus, 2009). White plague affected 
approximately ten percent of O. faveolata colonies and resulted in less 
than five percent tissue loss in all but two infected corals which 
experienced greater than five percent tissue loss (Brandt and McManus, 
2009).
    All sources of information are used to describe O. faveolata's 
susceptibility to ocean warming as follows. Orbicella faveolata is 
highly susceptible to elevated temperatures. In lab experiments, 
elevated temperatures resulted in misshapen embryos and differential 
gene expression in larvae that could indicate negative effects on 
larval development and survival. Bleaching susceptibility is generally 
high with 37 to 100 percent of O. faveolata colonies reported to bleach 
during several bleaching events. Chronic local stressors can exacerbate 
the effects of warming temperatures, which can result in slower 
recovery from bleaching, reduced calcification, and slower growth rates 
for several years following bleaching. Additionally, disease outbreaks 
affecting O. faveolata have been linked to elevated temperature as they 
have occurred after bleaching events. We conclude that O. faveolata is 
highly susceptible to elevated temperature.
    The SRR and SIR provided the following information on O. 
faveolata's susceptibility to acidification. A field study did not find 
any change in O. faveolata's calcification in field-

[[Page 53944]]

sampled colonies from the Florida Keys up through 1996.
    The public comments did not provide new or supplemental information 
on the susceptibility of O. faveolata to acidification. Supplemental 
information we found on the susceptibility of O. faveolata to 
acidification includes the following. In laboratory experiments, 
reproduction of O. faveolata was negatively impacted by increasing 
CO2, and impairment of fertilization was exacerbated at 
lower sperm concentrations (Albright, 2011b). Fertilization success was 
reduced by 25 percent at 529 [mu]atm (43 percent fertilization) and 40 
percent at 712 [mu]atm (34 percent fertilization) compared to controls 
at 435 [mu]atm (57 percent fertilization; Albright, 2011a). 
Additionally, growth rate of O. faveolata was reduced under lower pH 
conditions (7.6) compared to higher pH conditions (8.1) after 120 days 
of exposure (Hall et al., 2012).
    All sources of information are used to describe O. faveolata's 
susceptibility to acidification as follows. Laboratory studies indicate 
that O. faveolata is susceptible to ocean acidification both through 
reduced fertilization of gametes and reduced growth of colonies. Thus, 
we conclude that O. faveolata is highly susceptible to ocean 
acidification.
    The SRR and SIR did not provide any species-specific information on 
the susceptibility of O. faveolata to disease. The public comments also 
did not provide new or supplemental information on the susceptibility 
of O. faveolata to disease. Supplemental information we found on the 
susceptibility of O. faveolata to disease confirms the information on 
the Orbicella species complex and includes the following. Disease 
affected corals in Puerto Rico after the 2005 bleaching event, and O. 
faveolata was the species most affected (Bruckner and Hill, 2009). A 
1998 outbreak of white plague on three surveyed reefs in St. Lucia 
affected 19 percent of O. faveolata colonies, and O. faveolata was the 
species most affected (Nugues, 2002). Larger colonies in St. Lucia were 
more likely to get infected, but they were less likely to suffer 
complete mortality (Nugues, 2002). Tissue mortality of marked O. 
faveolata colonies was 51 percent, and no colonies showed regrowth 
during the 8 month study period (Nugues, 2002). Disease surveys 
conducted between August and December 1999 at 19 reef sites from six 
geographic areas across the wider Caribbean (Bermuda, Puerto Rico, 
Bonaire, Venezuela, Colombia, and Jamaica) revealed that O. faveolata 
showed the second highest incidence of disease at 4.7 to 10.4 percent 
across geographic locations (Weil et al., 2002).
    Surveys at five sites along the west coast of Dominica between 2000 
and 2002 revealed that O. faveolata was one of the species most 
susceptible to disease. Of the 12 species infected by white plague in 
2000, O. faveolata ranked second highest in disease prevalence (18.4 
percent of infected colonies were O. faveolata); it ranked third in 
2001 out of 14 species (12.7 percent) and second in 2002 out of 13 
species (18.8 percent). In addition, white plague infected the larger 
size classes of O. faveolata. Although only one colony experienced 
total colony mortality, O. faveolata had the highest amount of tissue 
loss in each year and in the three years combined (Borger and Steiner, 
2005).
    Yellow band disease in O. faveolata increased in abundance between 
1999 and 2004 on reefs near La Parguera and Desecheo and Mona Islands, 
Puerto Rico (Waddell, 2005). Yellow band disease mean lesion growth 
rates on O. faveolata in La Parguera, Puerto Rico had a significant 
positive correlation with mean yearly surface water temperatures 
between 1998 and 2010 (Burge et al., 2014). In Curacao colonies of O. 
faveolata infected with yellow band disease lost 90 percent of their 
tissue between 1997 and 2005 (Bruckner and Bruckner, 2006a). Only the 
unaffected parts of colonies continued to grow, and only the smallest 
lesions caused by disease healed (Bruckner and Bruckner, 2006a). 
Partial mortality was higher in 2005 (average of 40 percent) than in 
1998 (Bruckner and Bruckner, 2006a). Outbreaks of white plague occurred 
in 2001 and 2005 and infected O. faveolata and O. annularis with the 
highest frequency (Bruckner and Bruckner, 2006a).
    Yellow band disease significantly affects O. faveolata reproductive 
output. Fecundity of diseased lesions was significantly lower than 
transition and healthy-looking tissues on diseased colonies. Diseased 
lesions had 99 percent fewer eggs compared to un-diseased control 
colonies. Fecundity in transition areas was 24 percent less than 
healthy-looking areas of diseased colonies and was significantly lower 
(50 percent) than in un-diseased control colonies. Healthy-looking 
tissues of diseased colonies had 27 percent lower fecundity compared to 
un-diseased control colonies. Furthermore, in colonies that had 
recovered from disease, small tissue remnants (less than 100 cm\2\) had 
84 percent lower fecundity compared to un-diseased controls, and large 
tissue remnants (400 to 1000 cm\2\) had 64 percent lower fecundity 
compared to un-diseased controls (Weil et al., 2009).
    All sources of information are used to describe O. faveolata's 
susceptibility to disease as follows. Orbicella faveolata is often 
among the coral species with the highest disease prevalence and tissue 
loss. Outbreaks have been reported to affect ten to 19 percent of O. 
faveolata colonies, and yellow band disease and white plague have the 
greatest effect. Disease often affects larger colonies, and reported 
tissue loss due to disease ranges from five to 90 percent. 
Additionally, yellow band disease results in lower fecundity in 
diseased and recovered colonies of O. faveolata. Therefore, we conclude 
that O. faveolata is highly susceptible to disease.
    The SIR and SRR did not provide any species-specific information on 
the trophic effects of fishing on O. faveolata. The public comments did 
not provide new or supplemental information, and we did not find 
supplemental information on the trophic effects of fishing on O. 
faveolata. However, due to the level of reef fishing conducted in the 
Caribbean, coupled with Diadema die-off and lack of significant 
recovery, competition with algae can adversely affect coral 
recruitment. Thus, O. faveolata likely has some susceptibility to the 
trophic effects of fishing given its low recruitment rates. However, 
the available information does not support a more precise description 
of susceptibility to this threat.
    The SRR and SIR did not provide species-specific information on the 
susceptibility of O. faveolata to sedimentation, and the public 
comments did not provide new or supplemental information on its 
susceptibility to this threat. Supplemental information we found 
confirms the information on the susceptibility of the Orbicella species 
complex to sedimentation and includes the following. In St. Lucia, 
rates of partial mortality of O. annularis and O. faveolata were higher 
close to river mouths where sediments were deposited than they were 
farther from the river mouths, indicating the sensitivity of these two 
species to sedimentation (Nugues and Roberts, 2003).
    All sources of information are used to describe O. faveolata's 
susceptibility to sedimentation as follows. Sedimentation can cause 
partial mortality of O. faveolata, and genus-level information 
indicates that sedimentation negatively affects primary production, 
growth rates, calcification, colony size, and

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abundance. Therefore, we conclude that O. faveolata is highly 
susceptible to sedimentation.
    The SRR, SIR, and public comments did not provide information on 
the susceptibility of O. faveolata to nutrient enrichment, and we did 
not find any new or supplemental information on the susceptibility of 
O. faveolata to nutrient enrichment.
    All sources of information are used to describe O. faveolata's 
susceptibility to nutrient enrichment as follows. Although there is no 
species-specific information, the Orbicella species complex is 
susceptible to nutrient enrichment through reduced growth rates, 
lowered recruitment, and increased disease severity. Therefore, based 
on genus-level information, we conclude that O. faveolata is likely 
highly susceptible to nutrient enrichment.
    The SRR and SIR provided the following information on the 
susceptibility of O. faveolata to predation. Under laboratory 
conditions, black band disease was transmitted to healthy O. faveolata 
fragments in the presence of the butterflyfish Chaetodon capistratus 
but not in aquaria without the fish present, suggesting that the fish 
acts as a disease vector (Aeby and Santavy, 2006).
    The public comments did not provide new or supplemental information 
on the susceptibility of O. faveolata to predation. Supplemental 
information we found on the susceptibility of O. faveolata to predation 
includes the following. In surveys of the Florida Keys in 2012, two 
percent of O. faveolata colonies were affected by predation by the 
corallivorous snail C. abbreviata (Miller et al., 2013). Parrotfish 
consume O. annularis and O. faveolata more intensively than other coral 
species, but tissue regeneration capabilities appear to be high enough 
to counterbalance loss from predation (Mumby, 2009).
    All sources of information are used to describe O. faveolata's 
susceptibility to predation as follows. Orbicella faveolata is affected 
by a number of predators, but losses appear to be minimal. We conclude 
that O. faveolata has low susceptibility to predation.
    The SRR and SIR did not provide information on the effects of sea 
level rise on O. faveolata. The SRR described sea level rise as an 
overall low to medium threat for all coral species. The public comments 
did not provide new or supplemental information on O. faveolata's 
susceptibility to sea level rise, and we did not find any new or 
supplemental information. Thus, we conclude that O. faveolata has some 
susceptibility to sea level rise, but the available information does 
not support a more precise description of susceptibility to this 
threat.
    The SRR and SIR did not provide species-specific information on the 
susceptibility of O. faveolata to collection and trade, and the public 
comments did not provide new or supplemental information on its 
susceptibility to this threat. Supplemental information we found 
confirms the information in the SRR and SIR that collection and trade 
is not a significant threat for the Orbicella species complex. Over the 
last decade, collection and trade of this species has been primarily 
for scientific research rather than commercial purposes. Gross exports 
for collection and trade of O. faveolata between 2000 and 2012 averaged 
271 specimens (data available at http://trade.cites.org). We conclude 
that O. faveolata has low susceptibility to collection and trade.
Regulatory Mechanisms
    In the proposed rule, we relied on information from the Final 
Management Report for evaluating the existing regulatory mechanisms for 
controlling threats to all corals. However, we did not provide any 
species-specific information on the regulatory mechanism or 
conservation efforts for O. faveolata. Public comments were critical of 
that approach, and we therefore attempt to analyze regulatory 
mechanisms and conservation efforts on a species basis, where possible, 
in this final rule. Records confirm that O. faveolata occurs in five 
Atlantic ecoregions, and studies and observations have confirmed the 
presence of O. faveolata in an additional three ecoregions (Burman et 
al., 2012). These eight ecoregions encompass 26 kingdom's and 
countries' EEZs. The 26 kingdoms and countries are Antigua & Barbuda, 
Bahamas, Barbados, Belize, Colombia, Costa Rica, Cuba, Dominica, 
Dominican Republic, French Antilles, Grenada, Guatemala, Haiti, Kingdom 
of the Netherlands, Honduras, Jamaica, Mexico, Nicaragua, Panama, St. 
Kitts & Nevis, St. Lucia, St. Vincent & Grenadines, Trinidad and 
Tobago, United Kingdom (British Caribbean Territories and possibly 
Bermuda), United States (including U.S. Caribbean Territories), and 
Venezuela. The regulatory mechanisms relevant to O. faveolata, 
described first as a percentage of the above kingdoms and countries 
that utilize them to any degree, and second as the percentages of those 
kingdoms and countries whose regulatory mechanisms may be limited in 
scope, are as follows: General coral protection (31 percent with 12 
percent limited in scope), coral collection (50 percent with 27 percent 
limited in scope), pollution control (31 percent with 15 percent 
limited in scope), fishing regulations on reefs (73 percent with 50 
percent limited in scope), managing areas for protection and 
conservation (88 percent with 31 percent limited in scope). The most 
common regulatory mechanisms in place for O. faveolata are reef fishing 
regulations and area management for protection and conservation. 
However, half of the reef fishing regulations are limited in scope and 
may not provide substantial protection for the species. General coral 
protection and collection laws, along with pollution control laws, are 
much less common regulatory mechanisms for the management of O. 
faveolata.
Vulnerability to Extinction
    As explained above in the Risk Analyses section, a species' 
vulnerability to extinction results from the combination of its spatial 
and demographic traits, threat susceptibilities, and consideration of 
the baseline environment and future projections of threats. The SRR 
stated that the factors that increase the extinction risk for O. 
faveolata are its extremely low productivity (growth and recruitment), 
documented dramatic recent declines, and its restriction to the highly 
disturbed/degraded wider Caribbean region.
    Subsequent to the proposed rule, we received and gathered 
supplemental species- or genus-specific information, described above, 
that expands our knowledge regarding the species' abundance, 
distribution, and threat susceptibilities. We developed our assessment 
of the species' vulnerability to extinction using all the available 
information. As explained in the Risk Analyses section, our assessment 
in this final rule emphasizes the ability of the species' spatial and 
demographic traits to moderate or exacerbate its vulnerability to 
extinction, as opposed to the approach we used in the proposed rule, 
which emphasized the species' susceptibility to threats.
    The following characteristics of O. faveolata, in conjunction with 
the information described in the Corals and Coral Reefs section, Coral 
Habitat sub-section, and Threats Evaluation section above, affect its 
vulnerability to extinction currently and over the foreseeable future. 
The species has undergone major declines mostly due to warming-induced 
bleaching and disease. There is evidence of synergistic effects of 
threats for this species including disease outbreaks following

[[Page 53946]]

bleaching events and reduced thermal tolerance due to chronic local 
stressors stemming from land-based sources of pollution. Orbicella 
faveolata is highly susceptible to a number of threats, and cumulative 
effects of multiple threats have likely contributed to its decline and 
exacerbate vulnerability to extinction. Despite high declines, the 
species is still common and remains one of the most abundant species on 
Caribbean reefs. Its life history characteristics of large colony size 
and long life span have enabled it to remain relatively persistent 
despite slow growth and low recruitment rates, thus moderating 
vulnerability to extinction. However, the buffering capacity of these 
life history characteristics is expected to decrease as colonies shift 
to smaller size classes as has been observed in locations in its range. 
Its absolute population abundance has been estimated as at least tens 
of millions of colonies in each of several locations including the 
Florida Keys, Dry Tortugas, and the U.S. Virgin Islands and is higher 
than the estimate from these three locations due to the occurrence of 
the species in many other areas throughout its range. Despite the large 
number of islands and environments that are included in the species' 
range, geographic distribution in the highly disturbed Caribbean 
exacerbates vulnerability to extinction over the foreseeable future 
because O. faveolata is limited to an area with high, localized human 
impacts and predicted increasing threats. Its depth range of 0.5 to at 
least 40 m, possibly up to 90 m, moderates vulnerability to extinction 
over the foreseeable future because deeper areas of its range will 
usually have lower temperatures than surface waters, and acidification 
is generally predicted to accelerate most in waters that are deeper and 
cooler than those in which the species occurs. Orbicella faveolata 
occurs in most reef habitats, including both shallow and mesophotic 
reefs, which moderates vulnerability to extinction over the foreseeable 
future because the species occurs in numerous types of reef 
environments that are predicted, on local and regional scales, to 
experience highly variable thermal regimes and ocean chemistry at any 
given point in time. Its abundance, life history characteristics, and 
depth distribution, combined with spatial variability in ocean warming 
and acidification across the species' range, moderate vulnerability to 
extinction because the threats are non-uniform, and there will likely 
be a large number of colonies that are either not exposed or do not 
negatively respond to a threat at any given point in time.
Listing Determination
    In the proposed rule, using the determination tool formula 
approach, O. faveolata was proposed for listing as endangered because 
of: High vulnerability to ocean warming (E), disease (C), and ocean 
acidification (E); high vulnerability to sedimentation (A and E) and 
nutrient over-enrichment (A and E); decreasing trend in abundance (E); 
low relative recruitment rate (E); moderate overall distribution (based 
on narrow geographic distribution and wide depth distribution) (E); 
restriction to the Caribbean (E); and inadequacy of regulatory 
mechanisms (D).
    In this final rule, we changed the listing determination for O. 
faveolata from endangered to threatened. We made this determination 
based on a more species-specific and holistic approach, including 
consideration of the buffering capacity of this species' spatial and 
demographic traits, and the best available information above on O. 
faveolata's spatial structure, demography, threat susceptibilities, and 
management. This combination of factors indicates that O. faveolata is 
likely to become endangered throughout its range within the foreseeable 
future, and thus warrants listing as threatened at this time, because:
    (1) Orbicella faveolata is highly susceptible to ocean warming (ESA 
Factor E), disease (C), ocean acidification (E), sedimentation (A, E), 
and nutrients (A, E) and susceptible to trophic effects of fishing (A). 
These threats are expected to continue and increase into the future. In 
addition, the species is at heightened extinction risk due to 
inadequate existing regulatory mechanisms to address global threats 
(D);
    (2) Orbicella faveolata is geographically located in the highly 
disturbed Caribbean where localized human impacts are high and threats 
are predicted to increase as described in the Threats Evaluation 
section. A range constrained to this particular geographic area that is 
likely to experience severe and increasing threats indicates that a 
high proportion of the population of this species is likely to be 
exposed to those threats over the foreseeable future;
    (3) Orbicella faveolata has experienced substantial declines in 
abundance and percent cover over the past two decades; and
    (4) Orbicella faveolata's slow growth rate and low sexual 
recruitment limit its capacity for recovery from threat-induced 
mortality events throughout its range over the foreseeable future. 
Additionally, shifts to smaller size classes via fission and partial 
mortality of older, larger colonies, have reduced the buffering 
capacity of O. faveolata's life history strategy.
    The combination of these characteristics and future projections of 
threats indicates that the species is likely to be in danger of 
extinction within the foreseeable future throughout its range, and 
warrants listing as threatened at this time due to factors A, C, D, and 
E.
    The available information above on O. faveolata's spatial 
structure, demography, threat susceptibilities, and management also 
indicate that the species is not currently in danger of extinction and 
thus does not warrant listing as Endangered because:
    (1) While Orbicella faveolata's distribution within the Caribbean 
increases its risk of exposure to threats as described above, its known 
depth distribution is between 0.5 and 45 m, with occurrence by the 
complex as deep as 90 m, and its habitat includes various shallow and 
mesophotic reef environments. This moderates vulnerability to 
extinction currently because the species is not limited to one habitat 
type but occurs in numerous types of reef environments that will 
experience highly variable thermal regimes and ocean chemistry on local 
and regional scales at any given point in time, as described in more 
detail in the Coral Habitat and Threats Evaluation sections. There is 
no evidence to suggest that the species is so spatially fragmented that 
depensatory processes, environmental stochasticity, or the potential 
for catastrophic events currently pose a high risk to the survival of 
the species; and
    (2) Although O. faveolata's abundance has declined, it still has a 
common occurrence and remains one of the most dominant corals in the 
Caribbean. Its absolute abundance is at least tens of millions of 
colonies based on estimates from three locations. Absolute abundance is 
higher than estimates from these locations since it occurs in many 
other locations throughout its range. This absolute abundance allows 
for variation in the responses of individuals to threats to play a role 
in moderating vulnerability to extinction for the species to some 
degree, as described in more detail in the Corals and Coral Reefs 
section. There is no evidence of depensatory processes such as 
reproductive failure from low density of reproductive individuals and 
genetic processes such as inbreeding affecting this species. Thus, its 
absolute abundance indicates it is currently able to avoid high 
mortality from

[[Page 53947]]

environmental stochasticity, and mortality of a high proportion of its 
population from catastrophic events.
    The combination of these characteristics indicates that the species 
does not exhibit the characteristics of one that is currently in danger 
of extinction, as described previously in the Risk Analyses section, 
and thus does not warrant listing as endangered at this time.
    Range-wide, multitudes of conservation efforts are already broadly 
employed that are likely benefiting O. faveolata. However, considering 
the global scale of the most important threats to the species, and the 
ineffectiveness of conservation efforts at addressing the root cause of 
global threats (i.e., GHG emissions), we do not believe that any 
current conservation efforts or conservation efforts planned in the 
future will result in affecting the species' status to the point at 
which listing is not warranted.

Orbicella franksi

Introduction
    The SRR and SIR provided the following information on O. franksi's 
morphology. Orbicella franksi is distinguished by large, unevenly 
arrayed polyps that give the colony its characteristic irregular 
surface. Colony form is variable, and the skeleton is dense with poorly 
developed annual bands. Colony diameter can reach up to 5 m with a 
height of up to 2 m. The public comments did not provide new or 
supplemental information on O. franksi's morphology, and we did not 
find any new or supplemental information.
Spatial Information
    The SRR and SIR provided the following information on O. franksi's 
distribution, habitat, and depth range. Orbicella franksi is 
distributed in the western Atlantic and throughout the Caribbean Sea 
including in the Bahamas, Bermuda, and the Flower Garden Banks. 
Orbicella franksi tends to have a deeper distribution than the other 
two species in the Orbicella species complex.
    It occupies most reef environments and has been reported from water 
depths ranging from 5 to 50 m, with the species complex reported to 90 
m. Orbicella species are a common, often dominant, component of 
Caribbean mesophotic reefs, suggesting the potential for deep refugia 
for O. franksi.
    The public comments did not provide new or supplemental information 
on O. franksi's distribution, habitat, or depth range. We did not find 
new or supplemental information on O. franksi's habitat or depth range. 
Supplemental information we found on O. franksi's distribution includes 
the following. Veron (2014) confirmed the occurrence of O. franksi in 
six out of his 11 ecoregions in the western Atlantic and greater 
Caribbean known to contain corals and strongly predicted its presence 
in an additional three ecoregions (off Colombia/Venezuela, Cuba/Cayman 
Islands, and Jamaica). Other studies confirm the presence of O. franksi 
in three other ecoregions (Alcolado et al., 2010; Bayraktarov et al., 
2012; Bruckner, 2012c; Weil et al., 2002). The two ecoregions where O. 
franksi has not been found are off the coasts of Brazil and the 
southeastern U.S. north of southern Florida (Veron, 2014).
Demographic Information
    The SRR and SIR provided the following information on O. franksi's 
abundance and population trends. Orbicella franksi is reported as 
common.
    The public comments provided the following supplemental information 
on O. franksi's abundance and population trends. In surveys throughout 
the Florida Keys, O. franksi in 2005 ranked 26th most abundant out of 
47 coral species, 32nd out of 43 in 2009, and 33rd out of 40 in 2012. 
Extrapolated population estimates from stratified random surveys were 
8.0  3.5 million (SE) colonies in 2005, 0.3  
0.2 million (SE) colonies in 2009, and 0.4  0.4 million 
(SE) colonies in 2012. The authors note that differences in 
extrapolated abundance between years were more likely a function of 
sampling effort rather than an indication of population trends. In 
2005, the greatest proportions of colonies were in the smaller size 
classes of 10 to 20 cm and 20 to 30 cm. Partial colony mortality ranged 
from zero to approximately 73 percent and was generally higher in 
larger colonies (Miller et al., 2013).
    In the Dry Tortugas, Florida, O. franksi ranked fourth highest in 
abundance out of 43 coral species in 2006 and eighth out of 40 in 2008. 
Extrapolated population estimates were 79  19 million (SE) 
colonies in 2006 and 18.2  4.1 million (SE) colonies in 
2008. The authors note the difference in estimates between years was 
more likely a function of sampling effort rather than population 
decline. In the first year of the study (i.e., 2006), the greatest 
proportion of colonies were in the size class 20 to 30 cm with twice as 
many colonies as the next most numerous size class, and a fair number 
of colonies in the largest size class of greater than 90 cm. Partial 
colony mortality ranged from approximately ten to 55 percent. Two years 
later in 2008 no size class was found to dominate, and proportion of 
colonies in the medium to large size classes (60 to 90 cm) appeared to 
be less than in 2006. The number of colonies in the largest size class 
of greater than 90 cm remained consistent. Partial colony mortality 
ranged from approximately 15 to 75 percent (Miller et al., 2013).
    Supplemental information we found on O. franksi's abundance and 
population trends includes the following. In a 1995 survey of 16 reefs 
in the Florida Keys, O. franksi has the highest percent cover of all 
species (Murdoch and Aronson, 1999). In a survey of 31 sites in 
Dominica between 1999 and 2002, O. franksi was present in seven percent 
of the sites at less than one percent cover (Steiner, 2003). In 2003 on 
the east Flower Garden Bank, O. franksi comprised 46 percent of the 
76.5 percent coral cover on reefs 32 to 40 m in depth, and partial 
coral mortality due to bleaching, disease, and predation was rare in 
survey stations (Precht et al., 2005).
    Reported density is variable by location and habitat and is 
reported to range from 0.02 to 1.05 colonies per 10 m\2\. In surveys of 
1,176 sites in southeast Florida, the Dry Tortugas, and the Florida 
Keys between 2005 and 2010, density of O. franksi ranged between 0.04 
and 0.47 colonies per 10 m\2\ and was highest on the offshore patch 
reef and fore-reef habitats (Burman et al., 2012). In south Florida, 
density was highest in areas south of Miami at 0.44 colonies per 10 
m\2\ compared to 0.02 colonies per 10 m\2\ in Palm Beach and Broward 
Counties (Burman et al., 2012). Along the Florida reef tract from 
Martin County to the lower Florida Keys, density of O. franksi was 
approximately 0.9 colonies per 10 m\2\ (Wagner et al., 2010). On remote 
reefs off southwest Cuba, colony density was 0.083  0.17 
(SD) per 10 m transect on 38 reef-crest sites and 1.05  
1.02 colonies per 10 m transect on 30 reef-front sites (Alcolado et 
al., 2010). The number of O. franksi colonies in Cuba with partial 
colony mortality were far more frequent than those with no mortality 
across all size classes, except for one (i.e., less than 50 cm) that 
had similar frequency of colonies with and without partial mortality 
(Alcolado et al., 2010).
    In the U.S. Virgin Islands, O. franksi is the second most abundant 
species by percent cover at permanent monitoring stations. However, 
because the species complex, which is the most abundant by cover, was 
included as a category when individual Orbicella species could not

[[Page 53948]]

be identified with certainty, it is likely that O. franksi is the most 
abundant. Population estimates of O. franksi in the 49 km\2\ Red Hind 
Marine Conservation District are at least 34 million colonies (Smith, 
2013).
    Abundance in Curacao and Puerto Rico and appears to be stable over 
an eight to ten year period. In Curacao, abundance was stable between 
1997 and 2005, with partial mortality similar or less in 2005 compared 
to 1998 (Bruckner and Bruckner, 2006a). Abundance was also stable 
between 1998-2008 at nine sites off Mona and Desecheo Islands, Puerto 
Rico. In 1998, 4 percent of all corals at six sites surveyed off Mona 
Island were O. franksi colonies in 1998 and approximately five percent 
in 2008; at Desecheo Island, about two percent of all coral colonies 
were O. franksi in both 2000 and 2008 (Bruckner and Hill, 2009).
    On the other hand, colony size has decreased over the past several 
decades. A survey of 185 sites (2010 and 2011) in five countries 
(Bahamas, Bonaire, Cayman Islands, Puerto Rico, and St. Kitts and 
Nevis) reported the size of O. franksi and O. annularis colonies as 
significantly smaller than O. faveolata. The total mean partial 
mortality of O. franksi was 25 percent. Overall, the total live area 
occupied by O. franksi declined by a mean of 38 percent, and mean 
colony size declined from 1356 cm\2\ to 845 cm\2\. At the same time 
there was a 137 percent increase in small tissue remnants less than 500 
cm\2\, along with a decline in the proportion of large (1,500 to 30,000 
cm\2\), completely alive colonies. Mortality was attributed primarily 
to outbreaks of white plague and yellow band disease, which emerged as 
corals began recovering from mass bleaching events. This was followed 
by increased predation and removal of live tissue by damselfish to 
cultivate algal lawns (Bruckner, 2012a).
    All information on O. franksi's abundance and population trends can 
be summarized as follows. Based on population estimates, there are at 
least tens of millions of colonies present in both the Dry Tortugas and 
U.S. Virgin Islands. Absolute abundance is higher than the estimate 
from these two locations given the presence of this species in many 
other locations throughout its range. The frequency and extent of 
partial mortality, especially in larger colonies of O. franksi, appear 
to be high in some locations such as Florida and Cuba, though other 
locations like the Flower Garden Banks appear to have lower amounts of 
partial mortality. A decrease in O. franksi percent cover by 38 
percent, and a shift to smaller colony size across five countries, 
suggest that population decline has occurred in some areas; colony 
abundance appears to be stable in other areas. We conclude that while 
population decline has occurred, O. franksi is still common with the 
number of colonies at least in the tens of millions. Additionally, as 
discussed in the genus section, we conclude that the buffering capacity 
of O. franksi's life history strategy that has allowed it to remain 
abundant has been reduced by the recent population declines and amounts 
of partial mortality, particularly in large colonies.
Other Biological Information
    The SRR and SIR provided the following information on O. franksi's 
life history. The growth rate for O. franksi is reported to be slower, 
and spawning is reported to be about one to two hours earlier than O. 
annularis and O. faveolata.
    The public comments did not provide new or supplemental information 
on O. franksi's life history. Supplemental information we found on O. 
franksi's life history includes the following. Of 361 colonies of O. 
franksi tagged in Bocas del Toro, Panama, larger colonies were noted to 
spawn more frequently than smaller colonies between 2002 and 2009 
(Levitan et al., 2011). Darling et al. (2012) performed a biological 
trait-based analysis to categorize coral species into four life history 
strategies: Generalist, weedy, competitive, and stress-tolerant. The 
classifications were primarily separated by colony morphology, growth 
rate, and reproductive mode. Orbicella franksi was classified as a 
``generalist'' species, thus likely less vulnerable to environmental 
stress.
    The SRR and SIR provided the following other biological information 
on O. franksi. Low tissue biomass can render specific colonies of O. 
franksi susceptible to mortality from stress events, such as bleaching 
or disease. This suggests that differential mortality among 
individuals, species, and reefs from stress events such as bleaching or 
disease may be at least partially a function of differential colony 
biomass (indicating overall coral health) as opposed to genetic or 
physiologic differences among corals or their symbionts.
    In a 2010 cold-water event that affected south Florida, O. franksi 
ranked as the 14th most susceptible coral species out of 25 of the most 
abundant coral species. Average partial mortality was eight percent in 
surveys from Martin County to the lower Florida Keys after the 2010 
cold-water event compared to 0.4 percent average mortality during 
summer surveys between 2005 and 2009.
    The public comments did not provide new or supplemental biological 
information on O. franksi. Supplemental biological information we found 
on O. franksi includes the following. Of 351 O. franksi colonies 
observed to spawn at a site off Bocas del Toro, Panama, 324 were unique 
genotypes. Over 90 percent of O. franksi corals on this reef were the 
product of sexual reproduction, and 19 genetic individuals had 
asexually propagated colonies made up of two to four spatially adjacent 
ramets each. Individuals within a genotype spawned more synchronously 
than individuals of different genotypes. Additionally, within 5m, 
colonies nearby spawned more synchronously than farther spaced 
colonies, regardless of genotype. At distances greater than 5m, 
spawning was random between colonies (Levitan et al., 2011).
    In a study of symbiont composition of repeatedly sampled colonies 
of six species in the Bahamas and the Florida Keys (1998, and 2000 to 
2004), major changes in symbiont dominance over time were observed at 
certain Florida Keys reefs in O. annularis and O. franksi. Some 
colonies of O. annularis and O. franksi exhibited shifts in their 
associations attributed to recovery from the stresses of the 1997-98 
bleaching event. Most transitions in symbiont identity ended in 2002, 
three to five years after the 1997-98 bleaching event (Thornhill et 
al., 2006).
Susceptibility to Threats
    The threat susceptibility information from the SRR and SIR was 
interpreted in the proposed rule for O. franksi's vulnerability to 
threats as follows: High vulnerability to ocean warming, disease, 
acidification, sedimentation, and nutrient enrichment; moderate 
vulnerability to the trophic effects of fishing; and low vulnerability 
to sea level rise, predation, and collection and trade.
    The SRR and SIR did not provide species-specific information on the 
susceptibility of O. franksi to ocean warming. The public comments did 
not provide new or supplemental information on the susceptibility of O. 
franksi to ocean warming. Supplemental information we found on the 
susceptibility of O. franksi to ocean warming includes the following. A 
high percentage of O. franksi colonies experience bleaching during warm 
water temperature anomalies. Stratified random surveys on back-reefs 
and fore-reefs between one and 30 m depth off Puerto Rico (Mona and 
Desecho Islands,

[[Page 53949]]

La Parguera, Mayaguez, Boqueron, and Rincon) in 2005 and 2006 revealed 
severe bleaching in O. franksi with approximately 90 percent of 
colonies bleached (Waddell and Clarke, 2008). Surveys from 2005 to 2007 
along the Florida reef tract from Martin County to the lower Florida 
Keys indicated O. franksi had the tenth highest bleaching prevalence 
out of 30 species observed to bleach (Wagner et al., 2010). During a 
moderate bleaching event in Colombia in 2010, 88 percent of O. franksi 
bleached, and 12 percent paled at a site in Gayraca Bay (Bayraktarov et 
al., 2012). In 2011, 75 percent of O. franksi were dead and completely 
overgrown by algae (Bayraktarov et al., 2012). Based on samples of O. 
franksi and O. faveolata collected from the Mesoamerican Barrier Reef, 
calcification of these two species is projected to cease at 35 degrees 
C in this location in the absence of acidification (Carricart-Ganivet 
et al., 2012).
    All sources of information are used to describe O. franksi's 
susceptibility to ocean warming as follows. Available information 
indicates that O. franksi is highly susceptible to warming temperatures 
with a reported 88 to 90 percent bleaching frequency. Reported 
bleaching-related mortality from one study is high at 75 percent. There 
is indication that symbiont shuffling after bleaching in O. franksi. We 
conclude that O. franksi is highly susceptible to ocean warming.
    The SRR and SIR did not provide any species-specific information on 
the susceptibility of O. franksi to acidification, and the public 
comments did not provide new or supplemental information on its 
susceptibility to this threat. We did not find any new or supplemental 
information on the susceptibility of O. franksi to acidification. 
Although there is no species-specific information on the susceptibility 
of O. franksi to ocean acidification, genus information indicates that 
the species complex has reduced growth and fertilization success under 
acidic conditions. Thus, we conclude O. franksi likely has high 
susceptibility to ocean acidification.
    The SRR and SIR did not provide any species-specific information on 
the susceptibility of O. franksi to disease. The public comments did 
not provide new or supplemental information on the susceptibility of O. 
franksi to disease. Supplemental information we found on the 
susceptibility of O. franksi to disease includes the following. Disease 
surveys conducted between August and December 1999 at 19 reef sites 
from six geographic areas across the wider Caribbean (Bermuda, Puerto 
Rico, Bonaire, Venezuela, Colombia, and Jamaica) revealed that O. 
franksi had the third highest incidence of disease at 1.1 to 5.6 
percent across geographic locations (Weil et al., 2002). Between 1998 
and 2000, O. franksi was one of six coral species identified in the 
Virgin Islands as most susceptible to disease (Waddell, 2005). In 2004 
in Mexico, disease prevalence was highest in O. franksi with 41 percent 
of colonies infected, followed by 34 percent of O. annularis colonies 
and 31 percent of O. faveolata colonies (Ward et al., 2006). In Curacao 
colonies of O. franksi infected with yellow band disease lost an 
average of 30 percent of their tissue between 1997 and 2005, but some 
tagged colonies exhibited re-sheeting over disease lesions (Bruckner 
and Bruckner, 2006a).
    All sources of information are used to describe O. franksi's 
susceptibility to disease as follows. Orbicella franksi is often 
reported as among the species with the highest disease prevalence. 
Although there are few quantitative studies of the effects of disease 
on O. franksi, there is evidence that partial mortality can average 
about 25 to 30 percent and that disease can cause shifts to smaller 
size classes. Thus, we conclude that O. franksi is highly susceptible 
to disease.
    The SIR and SRR did not provide any species-specific information on 
the trophic effects of fishing on O. franksi. The public comments did 
not provide new or supplemental information, and we did not find new or 
supplemental information on the trophic effects of fishing on O. 
franksi. However, due to the level of reef fishing conducted in the 
Caribbean, coupled with Diadema die-off and lack of significant 
recovery, competition with algae can adversely affect coral 
recruitment. Thus, O. franksi likely has some susceptibility to the 
trophic effects of fishing given its low recruitment rates.
    The SRR, SIR, and public comments did not provide information on 
the susceptibility of O. franksi to sedimentation, and we did not find 
any new or supplemental information. All sources of information are 
used to describe O. franksi's susceptibility to sedimentation as 
follows. Genus information indicates sedimentation negatively affects 
primary production, growth rates, calcification, colony size, and 
abundance. Therefore, we conclude that O. franksi is highly susceptible 
to sedimentation.
    The SRR, SIR, and public comments do not provide information on the 
susceptibility of O. franksi to nutrient enrichment. Supplemental 
information we found on the susceptibility of O. franksi to nutrient 
enrichment includes the following. Field experiments indicate that 
nutrient enrichment significantly increases yellow band disease 
severity in O. annularis and O. franksi through increased tissue loss 
(Bruno et al., 2003).
    All sources of information are used to describe O. franksi's 
susceptibility to nutrient enrichment as follows. Genus level 
information indicates O. franksi is likely susceptible to nutrient 
enrichment through reduced growth rates and lower recruitment. 
Additionally, nutrient enrichment has been shown to increase the 
severity of yellow band disease in O. franksi. Thus, we conclude that 
O. franksi is highly susceptible to nutrient enrichment.
    The SRR and SIR do not provide species-specific information on the 
susceptibility of O. franksi to predation. Likewise, the public 
comments do not provide new or supplemental information on the 
susceptibility of O. franksi to predation. Supplemental information we 
found on the susceptibility of O. franksi to predation includes the 
following. Incidence of parrotfish grazing on the Belize barrier reef 
was second highest on O. franksi. However, in most habitats, the 
majority of Orbicella spp. showed little or no parrotfish grazing while 
only a few colonies were more heavily grazed, indicating low impact to 
the species overall (Rotjan, 2007).
    All sources of information are used to describe O. franksi's 
susceptibility to predation as follows. Genus-level information 
indicates O. franksi is affected by a number of predators, but both 
species-level and genus-level impacts appear to be minimal. We conclude 
that O. franksi has low susceptibility to predation.
    The SRR and SIR did not provide information on the effects of sea 
level rise on O. franksi. The SRR described sea level rise as an 
overall low to medium threat for all coral species. The public comments 
did not provide new or supplemental information on O. franksi's 
susceptibility to sea level rise, and we did not find any new or 
supplemental information. Thus, we conclude that O. franksi has some 
susceptibility to sea level rise, but the available information does 
not support a more precise description of susceptibility to this 
threat.
    The SRR and SIR do not provide species-specific information on the 
susceptibility of O. franksi to collection and trade, and the public 
comments do not provide new or supplemental information on its 
susceptibility to this threat. Supplemental information we found 
confirms the information in the

[[Page 53950]]

SRR and SIR that collection and trade is not a significant threat for 
the Orbicella species complex. Over the last decade, collection and 
trade of O. franksi has been primarily for scientific research rather 
than commercial purposes. Annual gross exports for collection and trade 
of O. franksi between 2000 and 2012 averaged 40 specimens (data 
available at http://trade.cites.org). Thus, we conclude that O. franksi 
has low susceptibility to collection and trade.
Regulatory Mechanisms
    In the proposed rule, we relied on information from the Final 
Management Report for evaluating the existing regulatory mechanisms for 
controlling threats to all corals. However, we did not provide any 
species-specific information on the regulatory mechanism or 
conservation efforts for O. franksi. Public comments were critical of 
that approach, and we therefore attempt to analyze regulatory 
mechanisms and conservation efforts on a species basis, where possible, 
in this final rule. Records confirm that O. franksi occurs in six 
Atlantic ecoregions, and studies have confirmed the presence of O. 
franksi in an additional three ecoregions. These nine ecoregions 
encompass 26 kingdoms' and countries' EEZs, and the 26 kingdoms and 
countries are Antigua & Barbuda, Bahamas, Barbados, Belize, Colombia, 
Costa Rica, Cuba, Dominica, Dominican Republic, French Antilles, 
Grenada, Guatemala, Haiti, Kingdom of the Netherlands, Honduras, 
Jamaica, Mexico, Nicaragua, Panama, St. Kitts & Nevis, St. Lucia, St. 
Vincent & Grenadines, Trinidad and Tobago, United Kingdom (British 
Caribbean Territories and Bermuda), United States (including U.S. 
Caribbean Territories), and Venezuela. The regulatory mechanisms 
relevant to O. franksi, described first as a percentage of the above 
kingdoms and countries that utilize them to any degree, and second as 
the percentage of those kingdoms and countries whose regulatory 
mechanisms may be limited in scope, are as follows: General coral 
protection (31 percent with 12 percent limited in scope), coral 
collection (50 percent with 27 percent limited in scope), pollution 
control (31 percent with 15 percent limited in scope), fishing 
regulations on reefs (73 percent with 50 percent limited in scope), 
managing areas for protection and conservation (88 percent with 31 
percent limited in scope). The most common regulatory mechanisms in 
place for O. franksi are reef fishing regulations and area management 
for protection and conservation. However, half of the reef fishing 
regulations are limited in scope and may not provide substantial 
protection for the species. General coral protection and collection 
laws, along with pollution control laws, are much less common 
regulatory mechanisms for the management of O. franksi.
Vulnerability to Extinction
    As explained above in the Risk Analyses section, a species' 
vulnerability to extinction results from the combination of its spatial 
and demographic traits, threat susceptibilities, and consideration of 
the baseline environment and future projections of threats. The SRR 
stated that the factors that increase the extinction risk for O. 
franksi are extremely low productivity (growth and recruitment), 
documented dramatic recent declines, and its restriction to the highly 
disturbed and degraded wider Caribbean region. All of these factors 
combined to yield a very high estimated extinction risk. It had a 
marginally lower risk estimate than the other two O. annularis complex 
species because of greater distribution in deep and mesophotic depth 
habitats, which are expected to experience lesser exposure to some 
surface-based threats.
    Subsequent to the proposed rule, we received and gathered 
supplemental species- or genus-specific information, described above, 
that expands our knowledge regarding the species' abundance, 
distribution, and threat susceptibilities. We developed our assessment 
of the species' vulnerability to extinction using all the available 
information. As explained in the Risk Analyses section, our assessment 
in this final rule emphasizes the ability of the species' spatial and 
demographic traits to moderate or exacerbate its vulnerability to 
extinction, as opposed to the approach we used in the proposed rule, 
which emphasized the species' susceptibility to threats.
    The following characteristics of O. franksi, in conjunction with 
the information described in the Corals and Coral Reefs section, Coral 
Habitat sub-section, and Threats Evaluation section above, affect its 
vulnerability to extinction currently and over the foreseeable future. 
The species has undergone declines most likely from disease and 
warming-induced bleaching. There is evidence of synergistic effects of 
threats for this species including increased disease severity with 
nutrient enrichment. Orbicella franksi is highly susceptible to a 
number of threats, and cumulative effects of multiple threats have 
likely contributed to its decline and exacerbate vulnerability to 
extinction. Despite declines, the species is still common and remains 
one of the most abundant species on Caribbean reefs. Its life history 
characteristics of large colony size and long life span have enabled it 
to remain relatively persistent despite slow growth and low recruitment 
rates, thus moderating vulnerability to extinction. However, the 
buffering capacity of these life history characteristics is expected to 
decrease as colonies shift to smaller size classes as has been observed 
in locations in its range. Its absolute population abundance has been 
estimated as at least tens of millions of colonies in both a portion of 
the U.S. Virgin Islands and the Dry Tortugas and is higher than the 
estimate from these two locations due to the occurrence of the species 
in many other areas throughout its range. Despite the large number of 
islands and environments that are included in the species' range, 
geographic distribution in the highly disturbed Caribbean exacerbates 
vulnerability to extinction over the foreseeable future because O. 
franksi is limited to an area with high, localized human impacts and 
predicted increasing threats. Its depth range of five to at least 50 m, 
possibly up to 90 m, moderates vulnerability to extinction over the 
foreseeable future because deeper areas of its range will usually have 
lower temperatures than surface waters, and acidification is generally 
predicted to accelerate most in waters that are deeper and cooler than 
those in which the species occurs. Orbicella franksi occurs in most 
reef habitats, including both shallow and mesophotic reefs, which 
moderates vulnerability to extinction over the foreseeable future 
because the species occurs in numerous types of reef environments that 
are predicted, on local and regional scales, to experience highly 
variable thermal regimes and ocean chemistry at any given point in 
time. Its abundance, life history characteristics, and depth 
distribution, combined with spatial variability in ocean warming and 
acidification across the species' range, moderate vulnerability to 
extinction because the threats are non-uniform, and there will likely 
be a large number of colonies that are either not exposed or do not 
negatively respond to a threat at any given point in time.
Listing Determination
    In the proposed rule, using the determination tool formula 
approach, O. franksi was proposed for listing as endangered because of: 
High vulnerability to ocean warming (E) disease (C), and ocean 
acidification (E); high vulnerability to sedimentation (A and E) and 
nutrient over-enrichment (A

[[Page 53951]]

and E); decreasing trend in abundance (E); low relative recruitment 
rate (E); moderate overall distribution (based on narrow geographic 
distribution and wide depth distribution (E); restriction to the 
Caribbean (E); and inadequacy of regulatory mechanisms (D).
    In this final rule, we changed the listing determination for O. 
franksi from endangered to threatened. We made this determination based 
on a more species-specific and holistic approach, including 
consideration of the buffering capacity of this species' spatial and 
demographic traits, and the best available information above on O. 
franksi's spatial structure, demography, threat susceptibilities, and 
management. This combination of factors indicates that O. franksi is 
likely to become endangered throughout its range within the foreseeable 
future, and thus warrants listing as threatened at this time, because:
    (1) Orbicella franksi is highly susceptible to ocean warming (ESA 
Factor E), disease (C), nutrients (A, E), ocean acidification (E), and 
sedimentation (A, E) and susceptible to trophic effects of fishing (A). 
These threats are expected to continue and increase into the future. In 
addition, the species is at heightened extinction risk due to 
inadequate existing regulatory mechanisms to address global threats 
(D);
    (2) Orbicella franksi is geographically located in the highly 
disturbed Caribbean where localized human impacts are high and threats 
are predicted to increase as described in the Threats Evaluation 
section. A range constrained to this particular geographic area that is 
likely to experience severe and increasing threats indicates that a 
high proportion of the population of this species is likely to be 
exposed to those threats over the foreseeable future;
    (3) Orbicella franksi has experienced a decline in benthic cover 
over the past two decades; and
    (4) Orbicella franksi's slow growth rate and low sexual recruitment 
limits its capacity for recovery from threat-induced mortality events 
throughout its range over the foreseeable future. Additionally, shifts 
to smaller size classes via fission and partial mortality of older, 
larger colonies, have reduced the buffering capacity of O. franksi's 
life history strategy.
    The combination of these characteristics and future projections of 
threats indicates that the species is likely to be in danger of 
extinction within the foreseeable future throughout its range, and 
warrants listing as threatened at this time due to factors A, C, D, and 
E.
    The available information above on O. franksi spatial structure, 
demography, threat susceptibilities, and management also indicate that 
the species is not currently in danger of extinction and thus does not 
warrant listing as Endangered because:
    (1) While Orbicella franksi's distribution within the Caribbean 
increases its risk of exposure to threats as described above, its known 
depth distribution is between 5 and 50 m, with occurrence by the 
species complex as deep as 90 m, and its habitat includes various 
shallow and mesophotic reef environments. This moderates vulnerability 
to extinction currently because the species is not limited to one 
habitat type but occurs in numerous types of reef environments that 
will experience highly variable thermal regimes and ocean chemistry on 
local and regional scales at any given point in time, as described in 
more detail in the Coral Habitat and Threats Evaluation sections. There 
is no evidence to suggest that the species is so spatially fragmented 
that depensatory processes, environmental stochasticity, or the 
potential for catastrophic events currently pose a high risk to the 
survival of the species;
    (2) Although O. franksi has declined in percent cover and colony 
size, there is evidence that population abundance has remained stable 
in some locations over a decadal time scale; and
    (3) Orbicella franksi has a common occurrence and remains one of 
the most dominant corals in the Caribbean. It has an absolute abundance 
of at least tens of millions of colonies based on estimates from two 
locations. Absolute abundance is higher than estimates from these 
locations since it occurs in many other locations throughout its range. 
This absolute abundance allows for variation in the responses of 
individuals to threats to play a role in moderating vulnerability to 
extinction for the species to some degree, as described in more detail 
in the Corals and Coral Reefs section. There is no evidence of 
depensatory processes such as reproductive failure from low density of 
reproductive individuals and genetic processes such as inbreeding 
affecting this species. Thus, its absolute abundance indicates it is 
currently able to avoid high mortality from environmental 
stochasticity, and mortality of a high proportion of its population 
from catastrophic events.
    The combination of these characteristics indicates that the species 
does not exhibit the characteristics of one that is currently in danger 
of extinction, as described previously in the Risk Analyses section and 
thus does not warrant listing as endangered at this time.
    Range-wide, multitudes of conservation efforts are already broadly 
employed that are likely benefiting O. franksi. However, considering 
the global scale of the most important threats to the species, and the 
ineffectiveness of conservation efforts at addressing the root cause of 
global threats (i.e., GHG emissions), we do not believe that any 
current conservation efforts or conservation efforts planned in the 
future will result in affecting the species' status to the point at 
which listing is not warranted.

Orbicella annularis

Introduction
    The SRR and SIR provided the following information on O. annularis' 
morphology. Orbicella annularis colonies grow in columns that exhibit 
rapid and regular upward growth. In contrast to the other two Orbicella 
species, margins on the sides of columns are typically dead. Live 
colony surfaces usually lack ridges or bumps. The public comments did 
not provide new or supplemental information on O. annularis' 
morphology, and we did not find any new or supplemental information.
Spatial Information
    The SRR and SIR provided the following information on the 
distribution, habitat and depth range of O. annularis. Orbicella 
annularis is common throughout the western Atlantic and greater 
Caribbean including the Flower Garden Banks but may be absent from 
Bermuda. Two personal communications were cited: one confirming its 
rarity in Bermuda, and the other stating O. annularis had not been seen 
in Bermuda. Orbicella annularis is reported from most reef environments 
in depths of 0.5 to 20 m. The Orbicella species complex is a common, 
often dominant component of Caribbean mesophotic reefs, suggesting the 
potential for deep refugia across a broader depth range, but O. 
annularis is generally described with a shallower distribution.
    The public comments did not provide new or supplemental information 
on O. annularis' distribution, habitat, or depth range. Supplemental 
information we found includes the following. Veron (2014) confirmed the 
occurrence of O. annularis in nine out of his 11 ecoregions in the 
western Atlantic and greater Caribbean known to contain corals, but 
indicated one of these ecoregions (Bermuda) has published records of 
occurrence that need further

[[Page 53952]]

investigation. Locke (2013) indicated early records of O. annularis in 
Bermuda may be incorrect since this species was historically 
undifferentiated from O. franksi and O. faveolata. The two ecoregions 
in which it is not found are off the coasts of Brazil and the 
southeastern U.S. north of southern Florida (Veron, 2014).
Demographic Information
    The SRR and SIR provided the following information on O. annularis' 
abundance and population trends. Orbicella annularis has been described 
as common overall. Demographic data collected in Puerto Rico over nine 
years straddling the 2005 bleaching event showed that population growth 
rates were stable in the pre-bleaching period (2001-2005) but declined 
one year after the bleaching event. Population growth rates declined 
even further two years after the bleaching event but returned to stasis 
the following year.
    The public comments provided the following supplemental information 
on O. annularis' abundance and population trends. In the Florida Keys, 
abundance of O. annularis ranked 30 out of 47 coral species in 2005, 13 
out of 43 in 2009, and 12 out of 40 in 2012. Extrapolated population 
estimates from stratified random samples were 5.6 million  
2.7 million (SE) in 2005, 11.5 million  4.5 million (SE) in 
2009, and 24.3 million  12.4 million (SE) in 2012. Size 
class distribution was somewhat variable between survey years, with a 
larger proportion of colonies in the smaller size classes in 2005 
compared to 2009 and 2012 and a greater proportion of colonies in the 
largest size class (>90 cm) in 2012 compared to 2005 and 2009. Partial 
colony mortality was lowest less than 10 cm (as low as approximately 5 
percent) up to approximately 70 percent in the larger size classes. In 
the Dry Tortugas, Florida, abundance of O. annularis ranked 41 out of 
43 in 2006 and 31 out of 40 in 2008. The extrapolated population 
estimate was 0.5 million  0.3 million (SE) colonies in 
2008. Differences in population estimates between years may be 
attributed to sampling effort rather than population trends (Miller et 
al., 2013).
    Supplemental information we found on O. annularis' abundance and 
population trends includes the following. In Utila, Honduras, O. 
annularis was present at 80 percent of sites surveyed between 1999 and 
2000 and was the second most common coral species (Afzal et al., 2001). 
In a survey of 31 sites in Dominica between 1999 and 2002, O. annularis 
was present at 20 percent of the sites at one to ten percent cover 
(Steiner, 2003).
    Colony density varies by habitat and location, and range from less 
than 0.1 to greater than one colony per 10 m\2\. In surveys of 1,176 
sites in southeast Florida, the Dry Tortugas, and the Florida Keys 
between 2005 and 2010, density of O. annularis ranged between 0.09 and 
0.84 colonies per 10 m\2\ and was highest on mid-channel reefs followed 
by inshore reefs, offshore patch reefs, and fore-reefs (Burman et al., 
2012). Along the east coast of Florida, density was highest in areas 
south of Miami (0.34 colonies per 10 m\2\) compared to Palm Beach and 
Broward Counties (0.04 colonies per 10 m\2\, Burman et al., 2012). In 
surveys between 2005 to 2007 along the Florida reef tract from Martin 
County to the lower Florida Keys, density of O. annularis was 
approximately 1.3 colonies per 10 m\2\ (Wagner et al., 2010). Off 
southwest Cuba on remote reefs, O. annularis density was 0.31  0.46 (SD) per 10 m transect on 38 reef-crest sites and 1.58 
 1.29 colonies per 10 m transect on 30 reef-front sites. 
Colonies with partial mortality were far more frequent than those with 
no partial mortality which only occurred in the size class less than 
100 cm (Alcolado et al., 2010).
    Population trends are available from a number of studies. In a 
study of sites inside and outside a marine protected area in Belize, O. 
annularis cover declined significantly over a ten year period (1998/99 
to 2008/09) (Huntington et al., 2011). In a study of ten sites inside 
and outside of a marine reserve in the Exuma Cays, Bahamas, cover of O. 
annularis increased between 2004 and 2007 inside the protected area and 
decreased outside the protected area (Mumby and Harborne, 2010). 
Between 1996 and 2006, O. annularis declined in cover by 37 percent in 
permanent monitoring stations in the Florida Keys (Waddell and Clarke, 
2008), and, cover of O. annularis in permanent monitoring stations 
between 1996 and 1998 on a reef in the upper Florida Keys declined 71 
percent (Porter et al., 2001).
    Orbicella annularis is the third most abundant coral by percent 
cover in permanent monitoring stations in the U.S. Virgin Islands. A 
decline of 60 percent was observed between 2001 and 2012 primarily due 
to bleaching in 2005. However, most of the mortality was partial 
mortality, and colony density in monitoring stations did not change 
(Smith, 2013).
    At nine sites off Mona and Desecheo Islands, Puerto Rico, no 
species extirpations were noted at any site over 10 years of monitoring 
between 1995 and 2008. However, O. faveolata and O. annularis sustained 
the largest losses with the number of colonies of O. annularis 
decreasing by 19 and 20 percent at Mona and Desecheo Islands, 
respectively. In 1998, eight percent of all corals at six sites 
surveyed off Mona Island were O. annularis colonies, dipping to 
approximately 6 percent in 2008. At Desecheo Island, 14 percent of all 
coral colonies were O. annularis in 2000 and 13 percent in 2008 
(Bruckner and Hill, 2009).
    Surveys of a degraded and a less degraded site in a marine 
protected area in Cartagena, Colombia, revealed that while large, old 
colonies of O. annularis were present, colonies had experienced high 
partial mortality that caused high fission rates and a dominance of 
small, non-reproductive ramets. Ramets that were non-reproductive or 
less fertile (less than 46 cm\2\) accounted for 72 percent and 55 
percent of the population at the surveyed sites, and only one percent 
and six percent of the ramets at the sites were large enough (200 
cm\2\) to be fully reproductive. In addition to the small ramet size, 
the lack of sexual recruitment led the authors to conclude that both 
populations were in decline, especially at the more degraded reef where 
mortality was higher and ramets were smaller, as individual colonies 
seemed to be growing old without being replaced (Alvarado-Chacon and 
Acosta, 2009).
    In a survey of 185 sites in five countries (Bahamas, Bonaire, 
Cayman Islands, Puerto Rico, and St. Kitts and Nevis) in 2010 to 2011, 
size of O. annularis and O. franksi colonies was significantly less 
than O. faveolata. Total mean partial mortality of O. annularis 
colonies at all sites was 40 percent. Overall, the total area occupied 
by live O. annularis declined by a mean of 51 percent, and mean colony 
size declined from 1927 cm\2\ to 939 cm\2\. There was a 211 percent 
increase in small tissue remnants less than 500 cm\2\, while the 
proportion of completely live large (1,500-30,000 cm\2\) colonies 
declined. Orbicella annularis colonies in Puerto Rico were much larger 
with large amounts of dead sections. In contrast, colonies in Bonaire 
were also large with greater amounts of live tissue. The presence of 
dead sections was attributed primarily to outbreaks of white plague and 
yellow band disease, which emerged as corals began recovering from mass 
bleaching events. This was followed by increased predation and removal 
of live tissue by damselfish algal lawns (Bruckner, 2012a).

[[Page 53953]]

    Hughes and Tanner (2000) documented the demographics of O. 
annularis in Jamaica from 1977 to 1993. At the beginning of the study, 
86 colonies were present within monitored stations. The number of 
colonies increased 40 to 42 percent between 1986 and 1987 due to 
fission (occurring at the same time as a decline in cover) and 
subsequently declined steadily to 40 colonies by 1993. Rates of 
survival, population growth, and recruitment declined over time, and 
the size structure became increasingly dominated by smaller size 
classes (Hughes and Tanner, 2000). Mortality increased sharply between 
1990 and 1993 due to the presence of smaller, more vulnerable colonies 
formed by partial mortality of larger colonies (Hughes, 1996). The 
persistence of large colonies had the greatest effect on population 
growth, and simulations indicated that the levels of recruitment needed 
to maintain population levels at 1977 levels increased sharply over 
time (Hughes and Tanner, 2000). Simulations with no sexual recruitment 
indicated that the population dynamics in the most recent period (1987 
to 1993) forecasted a population of zero within approximately 25 years. 
Simulation using the population dynamics observed between 1982 to 1987 
would result in a slower decline while the dynamics observed between 
1977 and 1982 would result in population growth (Hughes and Tanner, 
2000).
    Cover of O. annularis at Yawzi Point, St. John, U.S. Virgin Islands 
declined from 41 percent in 1988 to approximately 12 percent by 2003 
with a rapid decline beginning with the aftermath of Hurricane Hugo in 
1989 and continuing between 1994 and 1999 during a time of two 
hurricanes (1995) and a year of unusually high sea temperature (1998), 
and remaining statistically unchanged between 1999 and 2003. Colony 
abundances declined from 47 to 20 colonies per m\2\ between 1988 and 
2003, due mostly to the death and fission of medium to large colonies 
(>=151 cm\2\). Meanwhile, the population size class structure shifted 
between 1988 and 2003 to a higher proportion of smaller colonies in 
2003 (60 percent less than 50 cm\2\ in 1988 versus 70 percent in 2003) 
and lower proportion of large colonies (6 percent greater than 250 
cm\2\ in 1988 versus 3 percent in 2003). The changes in population size 
structure indicated a population decline coincident with the period of 
apparent stable coral cover. Population modeling forecasts the 1988 
size structure would not be reestablished by recruitment and a strong 
likelihood of extirpation of O. annularis at this site within 50 years 
(Edmunds and Elahi, 2007).
    Orbicella annularis colonies were monitored between 2001 and 2009 
at Culebra Island, Puerto Rico. The population was in demographic 
equilibrium (high rates of survival and stasis) before the 2005 
bleaching event but suffered a significant decline in growth rate 
(mortality and shrinkage) for two consecutive years after the bleaching 
event. Partial tissue mortality due to bleaching caused dramatic colony 
fragmentation that resulted in a population made up almost entirely of 
small colonies by 2007 (97 percent were less than 50 cm\2\). Three 
years after the bleaching event, the population stabilized at a number 
of colonies reduced by about half, with fewer medium to large size 
colonies and more smaller colonies (Hern[aacute]ndez-Pacheco et al., 
2011).
    All information on O. annularis' abundance and population trends 
can be summarized as follows. Historically, O. annularis was considered 
to be one of the most abundant species in the Caribbean (Weil and 
Knowlton, 1994). Percent cover has declined between 37 to 90 percent 
over the past several decades at reefs at Jamaica, Belize, Florida 
Keys, Bahamas, Bonaire, Cayman Islands, Curacao, Puerto Rico, U.S. 
Virgin Islands, and St. Kitts and Nevis. Based on population estimates, 
there are at least tens of millions of O. annularis colonies present in 
the Florida Keys and Dry Tortugas combined. Absolute abundance is 
higher than the estimate from these two locations given the presence of 
this species in many other locations throughout its range. Orbicella 
annularis remains common in occurrence. Abundance has decreased in some 
areas between 19 to 57 percent, and shifts to smaller size classes have 
occurred in locations such as Jamaica, Colombia, Bahamas, Bonaire, 
Cayman Islands, Puerto Rico, U.S. Virgin Islands, and St. Kitts and 
Nevis. At some reefs, a large proportion of the population is comprised 
of non-fertile or less-reproductive size classes. Several population 
projections indicate population decline in the future is likely at 
specific sites, and local extirpation is possible within 25 to 50 years 
at conditions of high mortality, low recruitment, and slow growth 
rates. We conclude that while substantial population decline has 
occurred in O. annularis, it is still common throughout the Caribbean 
and remains one of the dominant species numbering at least in the tens 
of millions of colonies. Additionally, as discussed in the genus 
section, we conclude that the buffering capacity of O. annularis' life 
history strategy that has allowed it to remain abundant has been 
reduced by the recent population declines and amounts of partial 
mortality, particularly in large colonies.
Other Biological Information
    The SRR and SIR provided the following information on O. annularis' 
life history. Orbicella annularis is reported to have slightly smaller 
egg size and potentially smaller size/age at first reproduction that 
the other two species of the Orbicella genus.
    The public comments did not provide new or supplemental information 
on the life history of O. annularis. Supplemental information we found 
on O. annularis' life history includes the following. The reported 
growth rate of O. annularis is 0.4 to 1.2 cm per year (Cruz-
Pi[ntilde][oacute]n et al., 2003; Tomascik, 1990). Darling et al. 
(2012) performed a biological trait-based analysis to categorize coral 
species into four life history strategies: Generalist, weedy, 
competitive, and stress-tolerant. The classifications were primarily 
separated by colony morphology, growth rate, and reproductive mode. 
Orbicella annularis was classified as a ``stress-tolerant'' species, 
thus likely less vulnerable to environmental stress.
    The SRR and SIR provided the following other biological information 
for O. annularis. Eight percent of O. annularis genotypes across three 
sites in Belize were clones. Low tissue biomass can render specific 
colonies of O. annularis susceptible to mortality from stress events, 
such as bleaching or disease. This suggests that differential mortality 
among individuals, species, and reefs from stress events such as 
bleaching or disease may be at least partially a function of 
differential colony biomass (indicating overall coral health) as 
opposed to genetic or physiologic differences among corals or their 
symbionts.
    In a 2010 cold-water event that affected south Florida, mortality 
of O. annularis was higher than any other coral species in surveys from 
Martin County to the lower Florida Keys. Average partial mortality was 
56 percent during the cold-water event compared to 0.3 percent from 
2005 to 2009. Surveys at a Florida Keys inshore patch reef, which 
experienced temperatures less than 18 degrees C for 11 days, revealed 
O. annularis was one of the most susceptible coral species with all 
colonies experiencing total colony mortality.
    The public comments did not provide new or supplemental biological 
information on O. annularis.

[[Page 53954]]

Supplemental biological information we found includes the following. Of 
117 colonies of O. annularis observed to spawn at a reef site off Bocas 
del Toro, Panama, there were 21 distinct genotypes, meaning that 82 
percent of the colonies were clones (Levitan et al., 2011). Individuals 
within a genotype spawned more synchronously than individuals of 
different genotypes. Colonies nearby spawned more synchronously 
regardless of genotype, out to about 5 m. When colonies were farther 
away, spawning was random.
    Of 137 O. annularis colonies sampled in Honduras, 118 were distinct 
genotypes, meaning that 14 percent of the colonies were clones. Over 90 
percent of genotypes were represented by a single colony, and 8 percent 
of the genotypes were represented by two or three colonies. One 
genotype had 14 colonies. Distance between clones ranged between 0.15 m 
to 6.94 m (Foster et al., 2007).
    Genetic sampling of 698 O. annularis colonies from 18 reefs within 
five countries in the Caribbean (Belize, Bahamas, Columbia, Curacao, 
and Honduras) revealed 466 distinct genotypes (approximately 33 percent 
clones). Genotypic diversity varied across the species' range from 
genetically diverse populations in Colombia, where every colony was 
unique, to genetically depauperate populations in Belize and Curacao, 
where a few genetic clones dominated. The contribution of clones to the 
local abundance of O. annularis increased in areas with greater 
hurricane frequency. Sites with higher genotypic diversity were 
dominated by larger colonies, and sites that experienced more frequent 
hurricanes were composed of smaller colonies than sites with less 
frequent hurricanes (Foster et al., 2013).
    Tissue samples of 1,424 O. annularis colonies at 26 reefs in 16 
regions of the Caribbean (Bahamas, Cuba, Dominican Republic, Puerto 
Rico, British Virgin Islands, Dominica, Barbados, Tobago, Venezuela, 
Netherlands Antilles, Colombia, Nicaragua, Jamaica, Cayman Islands, 
Belize, and Honduras) produced 943 distinct genotypes (34 percent 
clones). Three coarse population clusters of O. annularis were 
detected: eastern (Lesser Antilles, Venezuela, and Netherlands 
Antilles), western (Bahamas, Cuba, Belize, and Cayman Islands), and 
central (Jamaica, Honduras, Nicaragua, Colombia, Puerto Rico, British 
Virgin Islands, and Dominican Republic). No barrier to gene flow based 
on absolute geographic distance was apparent (Foster et al., 2012).
    In a study of symbiont composition of repeatedly sampled colonies 
of six species in the Bahamas and the Florida Keys in 1998 and 2000 to 
2004, major changes in symbiont dominance with time were observed in O. 
annularis and O. franksi at certain reefs in the Florida Keys. Some 
colonies of O. annularis and O. franksi exhibited shifts in their 
associations attributed to recovery from the stresses of the 1997-1998 
bleaching event. Most transitions in symbiont identity ended in 2002, 
three to five years after the 1997-98 bleaching event (Thornhill et 
al., 2006).
    All other biological information can be summarized as follows. 
Asexual fission and partial mortality can lead to multiple ramets. The 
percentage of unique genotypes is variable by location and is reported 
to range between 18 and 86 percent (14 to 82 percent are clones). 
Colonies in areas with higher disturbance from hurricanes tend to have 
more clonality. Genetic data indicate that there is some population 
structure in the eastern, central, and western Caribbean with 
population connectivity within areas but not across. Although O. 
annularis is still abundant, it may exhibit high clonality in some 
locations.
Susceptibility to Threats
    The threat susceptibility information from the SRR and SIR was 
interpreted in the proposed rule for O. annularis' vulnerabilities to 
threats as follows: High vulnerability to ocean warming, disease, 
acidification, sedimentation, and nutrient enrichment; moderate 
vulnerability to the trophic effects of fishing; and low vulnerability 
to sea level rise, predation, and collection and trade.
    The SRR and SIR provided the following information on the 
susceptibility of O. annularis to ocean warming. Simulation models 
using demographic data collected in Puerto Rico over nine years 
straddling the 2005 bleaching forecasted extinction of the population 
within 100 years at a bleaching frequency of once every five to ten 
years.
    The public comments did not provide new or supplemental information 
on the susceptibility of O. annularis to ocean warming. Supplemental 
information we found on the susceptibility of O. annularis to ocean 
warming includes the following. Surveys from 19 locations throughout 
the Caribbean indicated the bleaching event of 1995-96 was most 
extensive in the central and western Caribbean but only slight in the 
Lesser Antilles and Bermuda. Mortality of O. annularis from bleaching 
ranged from 2 to 30 percent at eight locations six months after the 
onset of bleaching (Alcolado et al., 2001).
    Eight of 15 colonies of O. annularis monitored in Jamaica from 1994 
to 1997 bleached in 1995. Bleaching affected polyp tissue depth, 
skeletal extension rate, reproduction, and density band formation, but 
the rate of recovery of each of these characteristics varied. Tissue 
depth recovered within 4 to 8 weeks after normal color returned, and 
growth rates returned to pre-bleaching levels once color and tissue 
depth returned. However, one year post bleaching, reproductive failure 
occurred in severely bleached colonies (bleached for 4 months and pale 
for an additional 3 months), and colonies that bleached mildly 
(bleached for 2 months with pale color for an additional 1 to 3 months) 
experienced reduced reproduction. Reproductive output of bleached 
colonies continued to be reduced two years after bleaching (Mendes and 
Woodley, 2002).
    Stratified random surveys on back-reefs and fore-reefs between one 
and 30 m depth off Puerto Rico (Mona and Desecho Islands, La Parguera, 
Mayaguez, Boqueron, and Rincon) in 2005 and 2006 revealed severe 
bleaching in O. annularis with greater than 95 percent of colonies 
bleached (Waddell and Clarke, 2008). Surveys from 2005 to 2007 along 
the Florida reef tract from Martin County to the lower Florida Keys 
indicated that O. annularis had the seventh highest bleaching 
prevalence out of 30 species observed (Wagner et al., 2010). During a 
2009 bleaching event on Little Cayman, of the ten coral species that 
bleached, O. annularis had the second highest bleaching prevalence with 
approximately 45 percent of colonies bleached (van Hooidonk et al., 
2012).
    Surveys at Culebra Island, Puerto Rico revealed extensive bleaching 
in 2005 with all of the O. annularis colonies in monitored transects 
bleached, and many of the surviving colonies remained pale in color 
after a year. Cover of O. annularis was reduced from 28 percent prior 
to the bleaching event in 2005 to 8 percent in 2009 (Hern[aacute]ndez-
Pacheco et al., 2011).
    In Barbados, the prevalence and abundance of the zooxanthellae 
Symbiodinium trenchi (D1a) increased in colonies of O. annularis in the 
weeks leading up to and during the 2005 bleaching event, and 
disproportionately dominated O. annularis colonies that did not bleach. 
In the 2-year period following the bleaching event, S. trenchi was 
displaced by other strains of Symbiodinium that were more competitive 
under less stressful conditions. The authors concluded that

[[Page 53955]]

it was unclear whether the rise and fall of S. trenchi was ecologically 
beneficial or whether its increased prevalence was an indicator of 
weakening coral health (LaJeunesse et al., 2009).
    Across the U.S. Virgin Islands, average bleaching of O. annularis 
was 66 percent, and paling was 16 percent, during the 2005 bleaching 
event. Disease prevalence of O. annularis was 5 percent after the 2005 
bleaching. In the milder 2010 bleaching event, 58 percent of O. 
annularis colonies bleached, and 28 percent of the colonies paled. No 
O. annularis colonies suffered total mortality, but percent cover 
decreased from the 2.5 percent cover in 2005 before bleaching to about 
one percent in 2010. There was a reduction in the proportion of larger 
sized colonies and an increase in the proportion of smaller sized 
colonies due to fission of larger colonies. The authors concluded that 
the susceptibility to disease increased the impact of bleaching, making 
O. annularis less tolerant overall to ocean warming (Smith et al., 
2013b).
    All sources of information are used to describe O. annularis' 
susceptibility to ocean warming as follows. Orbicella annularis is 
highly susceptible to bleaching with 45 to 100 percent of colonies 
observed to bleach. Reported mortality from bleaching ranges from two 
to 71 percent. Recovery after bleaching is slow with paled colonies 
observed for up to a year. Reproductive failure can occur a year after 
bleaching, and reduced reproduction has been observed two years post 
bleaching. There is indication that symbiont shuffling can occur prior 
to, during, and after bleaching events and result in bleaching 
resistance in individual colonies. We conclude that O. annularis is 
highly susceptible to ocean warming.
    The SRR and SIR did not provide any species-specific information on 
the susceptibility of O. annularis to acidification, and the public 
comments did not provide new or supplemental information on its 
susceptibility to this threat. We did not find any new or supplemental 
information on the susceptibility of O. annularis to acidification. 
Although there is no species-specific information on the susceptibility 
of O. annularis to ocean acidification, genus information indicates the 
species complex has reduced growth and fertilization success under 
acidic conditions. Thus, we conclude O. annularis likely has high 
susceptibility to ocean acidification.
    The SRR and SIR did not provide any species-specific information on 
the susceptibility of O. annularis to disease. The public comments did 
not provide new or supplemental information on the susceptibility of O. 
annularis to disease. Supplemental information we found on the 
susceptibility of O. annularis to disease confirms the information on 
the Orbicella species complex and includes the following. Surveys at 
five sites along the west coast of Dominica between 2000 and 2002 
revealed O. annularis was one of the species most susceptible to 
disease. Of the 12 species infected by white plague in 2000, O. 
annularis ranked third highest in disease prevalence (14.1 percent of 
infected colonies were O. annularis). It ranked second in 2001 out of 
14 species (20.3 percent) and third in 2002 out of 13 species (12.7 
percent). Although only one colony experienced total colony mortality, 
O. annularis had the third highest amount of tissue loss in the three 
years combined (11,717 cm\2\). Black band disease affected O. annularis 
in 2000 but not in any of the other survey years (Borger and Steiner, 
2005).
    In a 1998 outbreak of white plague in St. Lucia, three percent of 
O. annularis were affected, which was the lowest prevalence of disease 
of six species studied (Nugues, 2002). In surveys after the 2010 
bleaching event and the passage of a hurricane, 93 percent of diseased 
colonies (111 of 119 colonies) surveyed in radial transects in Brewers 
Bay, U.S. Virgin Islands were O. annularis (Brandt et al., 2013). 
Yellow band disease in O. annularis increased in prevalence between 
1999 and 2004 on reefs near La Parguera and Desecheo and Mona Islands, 
Puerto Rico (Waddell, 2005).
    Disease surveys conducted between August and December 1999 at 19 
reef sites from six geographic areas across the wider Caribbean 
(Bermuda, Puerto Rico, Bonaire, Venezuela, Colombia, and Jamaica) 
revealed that O. annularis showed the highest incidence of disease at 
5.5 to 12.6 percent across geographic locations. Yellow band disease 
showed higher incidences in Bonaire and Venezuela where a high 
proportion of recently dead ramets of O. annularis that most probably 
died from the disease were observed (Weil et al., 2002).
    In Curacao, colonies of O. annularis infected with yellow band 
disease lost 90 percent of their tissue between 1997 and 2005. Only the 
unaffected parts of colonies continued to grow, and only the smallest 
lesions healed. Partial mortality was higher in 2005 (average of 40 
percent) than in 1998. Outbreaks of white plague occurred in 2001 and 
2005 and infected O. faveolata and O. annularis with the highest 
frequency (Bruckner and Bruckner, 2006a).
    All sources of information are used to describe O. annularis' 
susceptibility to disease as follows. Most studies report O. annularis 
as among the species with the highest disease prevalence. Disease can 
cause extensive loss in coral cover, high levels of partial colony 
mortality, and changes in the relative proportions of smaller and 
larger colonies, particularly when outbreaks occur after bleaching 
events. Thus, we conclude that O. annularis is highly susceptible to 
disease.
    The SRR and SIR provided the following information on the 
susceptibility of O. annularis to the trophic effects of fishing. 
Interactions between O. annularis and four types of benthic algae 
(encrusting calcified red algae, fleshy brown macroalgae, upright 
calcareous green algae, and a mixed assemblage of turf algae) indicate 
that each alga exerts its own characteristic suite of effects on the 
coral holobiont, and that micro-scale dynamics have the potential to 
drive changes in reef community composition. Negative impacts spanned 
the range from micro-scale changes in microbial communities and oxygen 
drawdown to colony-scale effects such as damage to adjacent polyps and 
lowered fecundity of the adjacent colony. The public comments did not 
provide new or supplemental information on the susceptibility of O. 
annularis to the trophic effects of fishing, and we did not find any 
new or supplemental information.
    All sources of information are used to describe O. annularis' 
susceptibility to the trophic effects of fishing as follows. Due to the 
level of reef fishing conducted in the Caribbean, coupled with Diadema 
die-off and lack of significant recovery, competition with algae can 
adversely affect coral recruitment. In addition, competition with algae 
can lead to micro-scale to colony-level negative impacts to O. 
annularis. Thus, we conclude that O. annularis has some susceptibility 
to the trophic effects of fishing. The available information does not 
support a more precise description of susceptibility to this threat.
    The SRR and SIR did not provide species-specific information on the 
susceptibility of O. annularis to sedimentation, and the public 
comments did not provide new or supplemental information on its 
susceptibility to this threat. Supplemental information we found 
confirms the information on the susceptibility of the Orbicella species 
complex to sedimentation and includes the following. In St. Lucia, 
rates of partial mortality of O. annularis and O. faveolata were higher 
close to river

[[Page 53956]]

mouths where sediments were deposited than they were farther from the 
river mouths, indicating sensitivity of these two species to 
sedimentation (Nugues and Roberts, 2003). Additionally, at five study 
sites in Puerto Rico, the cover of O. annularis decreased significantly 
with a high content of terrigenous sediments (Torres and Morelock, 
2002).
    All sources of information are used to describe O. annularis' 
susceptibility to sedimentation as follows. Sedimentation can cause 
partial mortality and decreased coral cover of O. annularis. In 
addition, genus information indicates sedimentation negatively affects 
primary production, growth rates, calcification, colony size, and 
abundance. Therefore, we conclude that O. annularis has high 
susceptibility to sedimentation.
    The SRR, SIR, and public comments do not provide information on the 
susceptibility of O. annularis to nutrient enrichment. Supplemental 
information we found on the susceptibility of O. annularis to nutrient 
enrichment includes the following. Field experiments indicate that 
nutrient enrichment significantly increases yellow band disease 
severity in O. annularis and O. franksi through increased tissue loss 
(Bruno et al., 2003). In laboratory experiments, dissolved organic 
carbon caused significantly higher mortality of O. annularis after 30 
days of exposure compared to controls while nutrients (phosphate, 
nitrate, and ammonia) did not (Kline et al., 2006; Kuntz et al., 2005). 
Dissolved organic carbon levels that resulted in significantly higher 
mortality compared to controls were 12.5 mg per L glucose, and 25 mg 
per L lactose, starch, galactose, and glucose, which were all levels 
reported for impacted reefs (Kline et al., 2006; Kuntz et al., 2005).
    All sources of information are used to describe O. annularis' 
susceptibility to nutrient enrichment as follows. Elevated nutrients 
cause increased disease severity in O. annularis. Genus level 
information indicates elevated nutrients also cause reduced growth 
rates and lowered recruitment. Therefore, we conclude that O. annularis 
has high susceptibility to nutrients.
    The SRR and SIR do not provide species-specific information on the 
susceptibility of O. annularis to predation. Likewise, the public 
comments do not provide information on the susceptibility of O. 
annularis to predation. Supplemental information we found on the 
susceptibility of O. annularis to predation includes the following. 
Predation by the corallivorous snail C. abbreviata was present on 2.5 
percent of O. annularis colonies surveyed in the Florida Keys in 2012 
(Miller et al., 2013). Parrotfish consume O. annularis and O. faveolata 
more intensively than other coral species, but tissue regeneration 
capabilities appear to be high enough to counterbalance loss from 
predation (Mumby, 2009).
    All sources of information are used to describe O. annularis' 
susceptibility to predation as follows. Orbicella annularis is affected 
by a number of predators, but losses appear to be minimal. We conclude 
that O. annularis has low susceptibility to predation.
    The SRR and SIR did not provide information on the effects of sea 
level rise on O. annularis. The SRR described sea level rise as an 
overall low to medium threat for all coral species. The public comments 
did not provide new or supplemental information on O. annularis' 
susceptibility to sea level rise, and we did not find any new or 
supplemental information. Thus, we conclude that O. annularis has some 
susceptibility to sea level rise, but the available information does 
not support a more precise description of susceptibility to this 
threat.
    The SRR and SIR did not provide species-specific information on the 
susceptibility of O. annularis to collection and trade, and the public 
comments did not provide new or supplemental information on its 
susceptibility to this threat. Supplemental information we found 
confirms the information in the SRR and SIR that collection and trade 
is not a significant threat for the Orbicella species complex. Over the 
last decade, collection and trade of this species has been primarily 
for scientific research rather than commercial purposes. Annual gross 
exports for collection and trade of O. annularis between 2000 and 2012 
averaged 1,178 specimens (data available at http://trade.cites.org). 
Thus, we conclude that O. annularis has a low susceptibility to 
collection and trade.
Regulatory Mechanisms
    In the proposed rule, we relied on information from the Final 
Management Report for evaluating the existing regulatory mechanisms for 
controlling threats to all corals. However, we did not provide any 
species-specific information on the regulatory mechanism or 
conservation efforts for O. annularis. Public comments were critical of 
that approach, and we therefore attempt to analyze regulatory 
mechanisms and conservation efforts on a species basis, where possible, 
in this final rule. Records confirm that O. annularis occurs in nine 
Atlantic ecoregions that encompass 26 kingdom's and countries' EEZs. 
The 26 kingdoms and countries are Antigua & Barbuda, Bahamas, Barbados, 
Belize, Colombia, Costa Rica, Cuba, Dominica, Dominican Republic, 
French Antilles, Grenada, Guatemala, Haiti, Kingdom of the Netherlands, 
Honduras, Jamaica, Mexico, Nicaragua, Panama, St. Kitts & Nevis, St. 
Lucia, St. Vincent & Grenadines, Trinidad and Tobago, United Kingdom 
(British Caribbean Territories and possibly Bermuda), United States 
(including U.S. Caribbean Territories), and Venezuela. The regulatory 
mechanisms relevant to O. annularis, described first as a percentage of 
the above kingdoms and countries that utilize them to any degree, and 
second as the percentage of those kingdoms and countries whose 
regulatory mechanisms may be limited in scope, are as follows: General 
coral protection (31 percent with 12 percent limited in scope), coral 
collection (50 percent with 27 percent limited in scope), pollution 
control (31 percent with 15 percent limited in scope), fishing 
regulations on reefs (73 percent with 50 percent limited in scope), 
managing areas for protection and conservation (88 percent with 31 
percent limited in scope). The most common regulatory mechanisms in 
place for O. annularis are reef fishing regulations and area management 
for protection and conservation. However, half of the reef fishing 
regulations are limited in scope and may not provide substantial 
protection for the species. General coral protection and collection 
laws, along with pollution control laws, are much less common 
regulatory mechanisms for the management of O. annularis.
Vulnerability to Extinction
    As explained above in the Risk Analyses section, a species' 
vulnerability to extinction results from the combination of its spatial 
and demographic traits, threat susceptibilities, and consideration of 
the baseline environment and future projections of threats. The SRR 
stated that the factors that increase the extinction risk for O. 
annularis include very low productivity (growth and recruitment), 
documented dramatic declines in abundance, restriction to the degraded 
reefs of the wider Caribbean region, and preferential occurrence in 
shallow habitats (yielding potentially greater exposure to surface-
based threats.
    Subsequent to the proposed rule, we received and gathered 
supplemental species- or genus-specific information,

[[Page 53957]]

described above, that expands our knowledge regarding the species' 
abundance, distribution, and threat susceptibilities. We developed our 
assessment of the species' vulnerability to extinction using all the 
available information. As explained in the Risk Analyses section, our 
assessment in this final rule emphasizes the ability of the species' 
spatial and demographic traits to moderate or exacerbate its 
vulnerability to extinction, as opposed to the approach we used in the 
proposed rule, which emphasized the species' susceptibility to threats.
    The following characteristics of O. annularis, in conjunction with 
the information described in the Corals and Coral Reefs section, Coral 
Habitat sub-section, and Threats Evaluation section above, affect its 
vulnerability to extinction currently and over the foreseeable future. 
The species has undergone major declines mostly due to warming-induced 
bleaching and disease. Several population projections indicate 
population decline in the future is likely at specific sites and that 
local extirpation is possible within 25 to 50 years at conditions of 
high mortality, low recruitment, and slow growth rates. There is 
evidence of synergistic effects of threats for this species including 
disease outbreaks following bleaching events and increased disease 
severity with nutrient enrichment. Orbicella annularis is highly 
susceptible to a number of threats, and cumulative effects of multiple 
threats have likely contributed to its decline and exacerbate 
vulnerability to extinction. Despite high declines, the species is 
still common and remains one of the most abundant species on Caribbean 
reefs. Its life history characteristics of large colony size and long 
life span have enabled it to remain relatively persistent despite slow 
growth and low recruitment rates, thus moderating vulnerability to 
extinction. However, the buffering capacity of these life history 
characteristics is expected to decrease as colonies shift to smaller 
size classes as has been observed in locations in its range. Its 
absolute population abundance has been estimated as at least tens of 
millions of colonies in the Florida Keys and Dry Tortugas combined and 
is higher than the estimate from these two locations due to the 
occurrence of the species in many other areas throughout its range. 
Despite the large number of islands and environments that are included 
in the species' range, geographic distribution in the highly disturbed 
Caribbean exacerbates vulnerability to extinction over the foreseeable 
future because O. annularis is limited to an area with high, localized 
human impacts and predicted increasing threats. Orbicella annularis 
occurs in most reef habitats 0.5 to 20 m in depth which moderates 
vulnerability to extinction over the foreseeable future because the 
species occurs in numerous types of reef environments that are 
predicted, on local and regional scales, to experience highly variable 
thermal regimes and ocean chemistry at any given point in time. Its 
abundance and life history characteristics combined with spatial 
variability in ocean warming and acidification across the species' 
range, moderate vulnerability to extinction because the threats are 
non-uniform, and there will likely be a large number of colonies that 
are either not exposed or do not negatively respond to a threat at any 
given point in time.
Listing Determination
    In the proposed rule, using the determination tool formula 
approach, O. annularis was proposed for listing as endangered because 
of: High vulnerability to ocean warming (E) disease (C), and ocean 
acidification (E); high vulnerability to sedimentation (A and E) and 
nutrient over-enrichment (A and E); decreasing trend in abundance (E); 
low relative recruitment rate (E); narrow overall distribution (based 
on narrow geographic distribution and moderate depth distribution (E); 
restriction to the Caribbean; and inadequacy of regulatory mechanisms 
(D).
    In this final rule, we changed the listing determination for O. 
annularis from endangered to threatened. We made this determination 
based on a more species-specific and holistic approach, including 
consideration of the buffering capacity of this species' spatial and 
demographic traits, and the best available information above on O. 
annularis' spatial structure, demography, threat susceptibilities, and 
management. The combination of these factors indicates that O. 
annularis is likely to become endangered throughout its range within 
the foreseeable future, and thus warrants listing as threatened at this 
time, because:
    (1) Orbicella annularis is susceptible to ocean warming (ESA Factor 
E), disease (C), sedimentation (A, E), nutrients (A, E), and ocean 
acidification (E) and susceptible to trophic effects of fishing (A). 
These threats are expected to continue and increase into the future. In 
addition, the species is at heightened extinction risk due to 
inadequate existing regulatory mechanisms to address global threats 
(D);
    (2) Orbicella is geographically located in the highly disturbed 
Caribbean where localized human impacts are high and threats are 
predicted to increase as described in the Threats Evaluation section. A 
range constrained to this particular geographic area that is likely to 
experience severe and increasing threats indicates that a high 
proportion of the population of this species is likely to be exposed to 
those threats over the foreseeable future;
    (3) Orbicella annularis has undergone declines in abundance and 
percent cover over the past two decades;
    (4) Orbicella annularis' slow growth rate and low sexual 
recruitment limit its capacity for recovery from threat-induced 
mortality events throughout its range over the foreseeable future. 
Additionally, shifts to smaller size classes via fission and partial 
mortality of older, larger colonies, have reduced the buffering 
capacity of O. annularis' life history strategy; and
    (5) Several population projections and simulations predict 
continued population declines and local extirpation at specific sites 
within the foreseeable future.
    The combination of these characteristics and future projections of 
threats indicates that the species is likely to be in danger of 
extinction within the foreseeable future throughout its range, and 
warrants listing as threatened at this time due to factors A, C, D, and 
E.
    The available information above on O. annularis' spatial structure, 
demography, threat susceptibilities, and management also indicate that 
the species is not currently in danger of extinction and thus does not 
warrant listing as Endangered because:
    (1) While Orbicella annularis' distribution within the Caribbean 
increases its risk of exposure to threats as described above, its 
habitat includes most reef environments in water depths ranging from 
0.5 to 20 m. This moderates vulnerability to extinction currently 
because the species is not limited to one habitat type but occurs in 
numerous types of reef environments will experience highly variable 
thermal regimes and ocean chemistry on local and regional scales at any 
given point in time, as described in more detail in the Coral Habitat 
and Threats Evaluation sections. There is no evidence to suggest that 
the species is so spatially fragmented that depensatory processes, 
environmental stochasticity, or the potential for catastrophic events 
currently pose a high risk to the survival of the species;
    (2) Although O. annularis' abundance has declined, it still has a 
common

[[Page 53958]]

occurrence and remains one of the most dominant corals in the 
Caribbean. Its absolute abundance is at least tens of millions of 
colonies based on estimates from two locations. Absolute abundance is 
higher than estimates from these locations since it occurs in many 
other locations throughout its range. This absolute abundance allows 
for variation in the responses of individuals to threats to play a role 
in moderating vulnerability to extinction for the species to some 
degree, as described in more detail in the Corals and Coral Reefs 
section. There is no evidence of depensatory processes such as 
reproductive failure from low density of reproductive individuals and 
genetic processes such as inbreeding affecting this species. Thus, its 
absolute abundance indicates it is currently able to avoid high 
mortality from environmental stochasticity, and mortality of a high 
proportion of its population from catastrophic events; and
    (3) Some evidence shows that symbiont shuffling can occur prior to, 
during, and after bleaching events and result in bleaching resistance 
in individual colonies. This indicates O. annularis may have some 
buffering capacity against warming-induced bleaching.
    The combination of these characteristics indicates that the species 
does not exhibit the characteristics of one that is currently in danger 
of extinction, as described previously in the Risk Analyses section and 
thus does not warrant listing as endangered at this time.
    Range-wide, multitudes of conservation efforts are already broadly 
employed that are likely benefiting O. annularis. However, considering 
the global scale of the most important threats to the species, and the 
ineffectiveness of conservation efforts at addressing the root cause of 
global threats (i.e., GHG emissions), we do not believe that any 
current conservation efforts or conservation efforts planned in the 
future will result in affecting the species' status to the point at 
which listing is not warranted.

Genus Acropora (Caribbean)

    Acropora is the only genus considered in this rule that has species 
from both the Caribbean and the Indo-Pacific. Genus-level information 
for the Indo-Pacific species is described later under the section 
heading Genus Acropora (Indo-Pacific). Colonies in the Caribbean are 
all branching. There are over 300 nominal species in the genus 
Acropora, but in the Caribbean, there are only two species and one 
hybrid. Acropora cervicornis and A. palmata can interbreed to form the 
hybrid A. prolifera (Brainard et al., 2011). Acropora cervicornis shows 
genetic evidence of introgression or back-crossing with the hybrid A. 
prolifera while A. palmata does not (Brainard et al., 2011). The reason 
may be that A. palmata eggs are more resistant to fertilization in 
comparison to A. cervicornis eggs, as evidenced by an order of 
magnitude higher sperm needed to maximize conspecific fertilization, 
lower rates of heterospecific fertilization, and reduced viability 
after four hours (Fogarty et al., 2012c).
    Caribbean acroporiids are easily distinguishable and heavily 
studied. Therefore, this final rule does not provide an exhaustive 
discussion of the spatial, demographic, and threat vulnerabilities at 
the genus level. That information is described below for each species.

Acropora cervicornis

Introduction
    Acropora cervicornis is characterized by antler-like colonies with 
straight or slightly curved, cylindrical branches. The diameter of 
branches ranges from 0.25 to 5 cm (Lirman et al., 2010a), and linear 
branch growth rates have been reported to range between 3 and 11.5 cm 
per year (Acropora Biological Review Team, 2005). The species can exist 
as isolated branches, individual colonies up to about 1.5 m diameter, 
and thickets comprised of multiple colonies that are difficult to 
distinguish (Acropora Biological Review Team, 2005).
Spatial Information
    Information on A. cervicornis' distribution, habitat, and depth 
range that we considered in the proposed rule includes the following. 
Acropora cervicornis is distributed throughout the Caribbean, in the 
southwestern Gulf of Mexico, and in the western Atlantic. The fossil 
record indicates that during the Holocene, A. cervicornis was present 
as far north as Palm Beach County in southeast Florida (Lighty et al., 
1978), which is also the northern extent of its current distribution 
(Goldberg, 1973).
    Acropora cervicornis naturally occurs on spur and groove, bank 
reef, patch reef, and transitional reef habitats, as well as on 
limestone ridges, terraces, and hardbottom habitats (Cairns, 1982; 
Davis, 1982; Gilmore and Hall, 1976; Goldberg, 1973; Jaap, 1984; Miller 
et al., 2008; Wheaton and Jaap, 1988). It commonly grows in water 
ranging from five to 20 m in depth and has rarely been found to 60 m 
(Davis, 1982; Jaap, 1984; Jaap et al., 1989; Schuhmacher and Zibrowius, 
1985; Wheaton and Jaap, 1988). At the northern extent of its range, it 
grows in deeper water (16 to 30 m; Goldberg, 1973). Historically, 
staghorn coral was one of the primary constructors of mid-depth (10 to 
15 m) reef terraces in the western Caribbean, including Jamaica, the 
Cayman Islands, Belize, and some reefs along the eastern Yucatan 
peninsula (Adey, 1978). In the Florida Keys, A. cervicornis occurs in 
various habitats but is most prevalent on patch reefs as opposed to 
their former abundance in deeper fore-reef habitats (Miller et al., 
2008). There is no evidence of range constriction, though loss of A. 
cervicornis at the reef level has occurred (Acropora Biological Review 
Team, 2005).
    The public comments did not provide new or supplemental information 
on A. cervicornis' habitat or depth range. The public comments provided 
the following supplemental information on the distribution of A. 
cervicornis. Precht and Aronson (2004) postulate that coincident with 
climate warming, A. cervicornis only recently re-occupied its historic 
range after contracting to south of Miami, Florida during the late 
Holocene. They based this idea on the presence of large thickets off 
Ft. Lauderdale, Florida which were discovered in 1998 and had not been 
reported in the 1970s or 1980s (Precht and Aronson, 2004). However, 
because the presence of A. cervicornis in Palm Beach County, north of 
Ft. Lauderdale, was reported in the early 1970s (though no thicket 
formation was reported; Goldberg, 1973), there is uncertainty 
associated with whether these thickets were present prior to their 
discovery or if they recently appeared coincident with warming.
    We did not find any new or supplemental information on habitat or 
depth range. Supplemental information we found on A. cervicornis' 
distribution is consistent with information considered in the proposed 
rule and includes the following. Veron (2014) confirms the presence of 
A. cervicornis in seven out of a potential 11 ecoregions in the western 
Atlantic and greater Caribbean that are known to contain corals. The 
four ecoregions in which it is not found are the Flower Garden Banks 
and off the coasts of Bermuda, Brazil, and the southeast U.S. north of 
south Florida. The proportion of reefs with A. cervicornis present 
decreased dramatically after the Caribbean-wide mass mortality in the 
1970s and 1980s, indicating the spatial structure of the species has 
been affected by extirpation from many localized areas throughout its 
range (Jackson et al., 2014).

[[Page 53959]]

Demographic Information
    Information on A. cervicornis' abundance and population trends that 
we considered in the proposed rule includes the following. Acropora 
cervicornis has been described as sometimes common (Veron, 2000) and 
uncommon (Carpenter et al., 2008). Acropora cervicornis historically 
was one of the dominant species on most Caribbean reefs, forming large, 
monotypic thickets and giving rise to the nominal distinct zone in 
classical descriptions of Caribbean reef morphology (Goreau, 1959). 
Massive, Caribbean-wide mortality, apparently primarily from white band 
disease (Aronson and Precht, 2001), spread throughout the Caribbean in 
the mid-1970s to mid-1980s and precipitated widespread and radical 
changes in reef community structure (Brainard et al., 2011). In 
addition, continuing coral mortality from periodic acute events such as 
hurricanes, disease outbreaks, and mass bleaching events has added to 
the decline of A. cervicornis (Brainard et al., 2011). In locations 
where quantitative data are available (Florida, Jamaica, U.S. Virgin 
Islands, Belize), there was a reduction of approximately 92 to greater 
than 97 percent between the 1970s and early 2000s (Acropora Biological 
Review Team, 2005).
    Fossil evidence from the Dominican Republic indicates that Holocene 
A. cervicornis was capable of thriving for thousands of years under 
highly variable temperature and salinity conditions and suggests that 
the recent decline in A. cervicornis is anomalous (Greer et al., 2009). 
Additional fossil evidence from Belize indicates that the recent 
decline of A. cervicornis is without precedent during the late Holocene 
(Aronson and Precht, 2001). In contrast, two 500 year gaps in the 
fossil record, around 3 and 4.5 thousand years ago where dated A. 
cervicornis fragments were not observed in samples from the Florida 
Keys, suggests that the recent decline may not be without precedent 
(Shinn et al., 2003). However, this study was based on radiocarbon 
dating of A. cervicornis fragments, for which the time of transport and 
deposition are not known, so there is uncertainty of whether these gaps 
represent the absence of A. cervicornis or variable storm depositional 
history (Shinn et al., 2003).
    Since the 2006 listing of A. cervicornis as threatened, continued 
population declines have occurred in some locations with certain 
populations of both species decreasing up to an additional 50 percent 
or more (Colella et al., 2012; Lundgren and Hillis-Starr, 2008; Muller 
et al., 2008; Rogers and Muller, 2012; Williams et al., 2008).
    Public comments provided the following supplemental information on 
A. cervicornis' abundance and population trends. There are some small 
pockets of remnant robust populations such as in southeast Florida 
(Vargas-Angel et al., 2003), Honduras (Keck et al., 2005; Riegl et al., 
2009), and Dominican Republic (Lirman et al., 2010a). Additionally, 
Lidz and Zawada (2013) observed 400 colonies of A. cervicornis along 
70.2 km of transects near Pulaski Shoal in the Dry Tortugas where the 
species had not been seen since the cold water die-off of the 1970s. 
Cover of A. cervicornis increased on a Jamaican reef from 0.6 percent 
in 1995 to 10.5 percent in 2004 (Idjadi et al., 2006).
    Riegl et al. (2009) monitored A. cervicornis in photo plots on the 
fringing reef near Roatan, Honduras from 1996 to 2005. Acropora 
cervicornis cover was 0.42 percent in 1996, declined to 0.14 percent in 
1999 after the Caribbean bleaching event in 1998 and mortality from 
run-off associated with a Category 5 hurricane, and decreased further 
to 0.09 percent in 2005. Acropora cervicornis colony frequency 
decreased 71 percent between 1997 and 1999. In sharp contrast, offshore 
banks near Roatan had dense thickets of A. cervicornis with 31 percent 
cover in photo-quadrats in 2005 and appeared to survive the 1998 
bleaching event and hurricane, most likely due to bathymetric 
separation from land and greater flushing. Modeling showed that under 
undisturbed conditions, retention of the dense A. cervicornis stands on 
the banks off Roatan is likely with a possible increased shift towards 
dominance by other coral species. However, the authors note that 
because their data and the literature seem to point to extrinsic 
factors as driving the decline of A. cervicornis, it is unclear what 
the future may hold for this dense population (Riegl et al., 2009).
    Miller et al. (2013) extrapolated population abundance of A. 
cervicornis in the Florida Keys and Dry Tortugas from stratified random 
samples across habitat types. Population estimates of A. cervicornis in 
the Florida Keys were 10.2  4.6 (SE) million colonies in 
2005, 6.9  2.4 (SE) million colonies in 2007, and 10.0 
 3.1 (SE) million colonies in 2012. In the Dry Tortugas 
population estimates were 0.4  0.4 (SE) million colonies in 
2006 and 3.5  2.9 (SE) million colonies in 2008, though the 
authors note their sampling scheme in the Dry Tortugas was not 
optimized for A. cervicornis. Because these population estimates were 
based on random sampling, differences in abundance estimates between 
years may be more likely a function of sampling effort rather than 
population trends. In both the Florida Keys and Dry Tortugas, most of 
the population was dominated by small colonies less than 30 cm 
diameter. In the Florida Keys, partial mortality was highest in 2005, 
with up to 80 percent mortality observed, and lowest in 2007 with a 
maximum of 30 percent. In 2012, partial mortality ranged from 20 to 50 
percent across most size classes.
    Supplemental information we found on A. cervicornis' abundance and 
population trends includes the following. Acropora cervicornis was 
observed in 21 out of 301 stations between 2011 and 2013 in stratified 
random surveys designed to detect Acropora colonies along the south, 
southeast, southwest, and west coasts of Puerto Rico, and it was 
observed at an additional 16 sites outside of the surveyed area 
(Garc[iacute]a Sais et al., 2013). The largest colony was 60 cm, and 
density ranged from 1 to 10 colonies per 15 m\2\ (Garc[iacute]a Sais et 
al., 2013).
    While cover of A. cervicornis increased from 0.6 percent in 1995 to 
10.5 percent in 2004 (Idjadi et al., 2006) and 44 percent in 2005 on a 
Jamaican reef, it collapsed after the 2005 bleaching event and 
subsequent disease to less than 0.5 percent in 2006 (Quinn and Kojis 
2008). A cold water die-off in the Florida Keys in January 2010 
resulted in the complete mortality of all A. cervicornis colonies at 45 
of the 74 reefs surveyed (61 percent), spanning the lower to upper 
Florida Keys (Schopmeyer et al., 2012). Walker et al. (2012) report 
increasing size of two thickets (expansion of up to 7.5 times the 
original size of one of the thickets) monitored off southeast Florida 
and also noted that cover within monitored plots concurrently decreased 
by about 50 percent, highlighting the dynamic nature of A. cervicornis 
distribution via fragmentation and re-attachment.
    New information we found on population trends includes the 
following. A report on the status and trends of Caribbean corals over 
the last century indicates that cover of A. cervicornis has remained 
relatively stable (though much reduced) throughout the region since the 
large mortality events of the 1970s and 1980s. The frequency of reefs 
at which A. cervicornis was described as the dominant coral has 
remained stable. The number of reefs with A. cervicornis present 
declined during the 1980s (from approximately 50 to 30 percent of 
reefs), remained relatively stable through the

[[Page 53960]]

1990s, and decreased to approximately 20 percent of the reefs 2000-
2004, and approximately 10 percent 2005-2011 (Jackson et al., 2014).
    We summarize all sources of information on A. cervicornis' 
abundance and population trends as follows. Based on population 
estimates, there are at least tens of millions of colonies present in 
the Florida Keys and Dry Tortugas combined. Absolute abundance is 
higher than the estimate from these two locations given the presence of 
this species in many other locations throughout its range. The 
effective population size is smaller than indicated by abundance 
estimates due to the tendency for asexual reproduction. There is no 
evidence of range constriction or extirpation at the island level. 
However the species is absent at the reef level. Populations appear to 
consist mostly of isolated colonies or small groups of colonies 
compared to the vast thickets once prominent throughout its range, with 
thickets still a prominent feature at only a handful of known 
locations. Across the Caribbean, percent cover appears to have remained 
relatively stable since the population crash in the 1980s. Frequency of 
occurrence has decreased since the 1980s. There are examples of 
increasing trends in some locations (Dry Tortugas and southeast 
Florida), but not over larger spatial scales or longer time frames. 
Population model projections from Honduras at one of the only known-
remaining thickets indicate the retention of this dense stand under 
undisturbed conditions. If refuge populations are able to persist, it 
is unclear whether they would be able to repopulate nearby reefs as 
observed sexual recruitment is low. Thus, we conclude that the species 
has undergone substantial population decline and decreases in the 
extent of occurrence throughout its range. Percent benthic cover and 
proportion of reefs where A. cervicornis is dominant have remained 
stable since the mid-1980s and since the listing of the species as 
threatened in 2006. We also conclude that population abundance is at 
least tens of millions of colonies, but likely to decrease in the 
future with increasing threats.
Other Biological Information
    Information on A. cervicornis' life history that we considered in 
the proposed rule includes the following. Acropora cervicornis is a 
hermaphroditic broadcast spawning species. The spawning season occurs 
several nights after the full moon in July, August, or September, but 
may be split over the course of more than one lunar cycle (Szmant, 
1986; Vargas-Angel et al., 2006). The estimated size at sexual maturity 
is 17 cm branch length, and large colonies produce proportionally more 
gametes than small colonies (Soong and Lang, 1992). Basal and branch 
tip tissue is not fertile (Soong and Lang, 1992). Sexual recruitment 
rates are low, and this species is generally not observed in coral 
settlement studies. However, laboratory studies have found that certain 
species of crustose-coralline algae facilitate larval settlement and 
post-settlement survival (Ritson-Williams et al., 2010).
    Reproduction occurs primarily through asexual fragmentation that 
produces multiple colonies that are genetically identical (Tunnicliffe, 
1981). The combination of branching morphology, asexual fragmentation, 
and fast growth rates can lead to persistence of large areas dominated 
by A. cervicornis.
    The public comments did not provide new or supplemental information 
on A. cervicornis' life history. Supplemental information we found on 
life history includes the following. Darling et al. (2012) performed a 
biological trait-based analysis to categorize coral species into four 
life history strategies: Generalist, weedy, competitive, and stress-
tolerant. The classifications were primarily separated by colony 
morphology, growth rate, and reproductive mode. Acropora cervicornis 
was classified as a ``competitive'' species, thus likely more 
vulnerable to environmental stress.
    All information on A. cervicornis' life history can be summarized 
as follows. The combination of rapid skeletal growth rates and frequent 
asexual reproduction by fragmentation can enable effective competition 
and can facilitate potential recovery from disturbances when 
environmental conditions permit. However, low sexual reproduction can 
lead to reduced genetic diversity and limits the capacity to repopulate 
sites.
    Other biological information on A. cervicornis that we considered 
in the proposed rule includes the following. Vollmer and Palumbi (2007) 
examined 22 populations of A. cervicornis from nine regions in the 
Caribbean (Panama, Belize, Mexico, Florida, Bahamas, Turks and Caicos, 
Jamaica, Puerto Rico, and Cura[ccedil]ao) and concluded that 
populations greater than 500 km apart are genetically differentiated 
with low gene flow across the greater Caribbean. Fine-scale genetic 
differences have been detected at reefs separated by as little as 2 km, 
suggesting that gene flow in A. cervicornis may not occur at much 
smaller spatial scales (Garcia Reyes and Schizas, 2010; Vollmer and 
Palumbi, 2007). This fine-scale population structure was greater when 
considering genes of A. palmata introgressed in A. cervicornis due to 
back-crossing of the hybrid A. prolifera with A. cervicornis (Garcia 
Reyes and Schizas, 2010; Vollmer and Palumbi, 2007).
    Populations in Florida and Honduras are genetically distinct from 
each other and other populations in the U.S. Virgin Islands, Puerto 
Rico, Bahamas, and Navassa (Baums et al., 2010), indicating little to 
no larval connectivity. However, some potential connectivity between 
the U.S. Virgin Islands and Puerto Rico was detected and also between 
Navassa and the Bahamas (Baums et al., 2010).
    Florida populations of A. cervicornis have high levels of both 
genetic diversity and connectivity, with evidence suggesting the 
western Caribbean has historically been the source of genetic variation 
for Florida (Hemond and Vollmer, 2010). Colonies of A. cervicornis in 
Florida mostly harbored zooxanthellae Clade A, but colonies from 
inshore and mid-channel reefs, which experience higher sedimentation 
and temperature fluctuations than reefs further offshore, had a higher 
prominence of Clades C and D, revealing the influence of habitat on 
zooxanthellae associations (Baums et al., 2010).
    The public comments did not provide new or supplemental biological 
information on A. cervicornis, and we did not find any new or 
supplemental biological information. All information on A. cervicornis' 
biology can be summarized as follows. Connectivity over distances of 
greater than 500 km is limited, and there is evidence of restricted 
gene flow over much smaller spatial scales. Genetic diversity appears 
to be relatively high in some areas like the Florida Keys.
Susceptibility to Threats
    Information on threat susceptibilities was interpreted in the 
proposed rule for A. cervicornis' vulnerabilities to threats as 
follows: High vulnerability to ocean warming, disease, acidification, 
sedimentation, and nutrient enrichment; moderate vulnerability to the 
trophic effects of fishing and predation; and low vulnerability to sea 
level rise and collection and trade.
    Information on A. cervicornis' susceptibility to disease that we 
considered in the proposed rule includes the following. Disease is 
believed to be the primary cause of the region-wide decline of A. 
cervicornis beginning in the late 1970s (Aronson and Precht, 2001) and 
continues to have a large impact on the species. White band disease is 
generally associated

[[Page 53961]]

with the majority of disease-related mortalities, but several other 
diseases affect A. cervicornis. Ritchie and Smith (1995; 1998) 
described white band disease type II which is linked with a bacterial 
infection by Vibrio carchariae (Ritchie and Smith, 1998), also referred 
to as V. charchariae and V. harveyi (Gil-Agudelo et al., 2006). 
Williams and Miller (2005) reported an outbreak of a transmissible 
disease that caused rapid tissue loss on A. cervicornis in the Florida 
Keys in 2003. The disease manifested as irregular, multifocal tissue 
lesions with apparently healthy tissue remaining in between, a 
description similar to A. palmata afflicted with white pox. 
Additionally ciliate infections have been reported by Croquer et al. 
(2006) at several locations in the Caribbean.
    Few studies follow the progression of disease in individual 
colonies over time, but there are reports of instantaneous levels of 
disease at various locations. The Acropora Biological Review Team 
(2005) reported that in the 1997 to 2000 AGRRA surveys, at least 6 
percent of A. cervicornis colonies were diseased, with greater 
prevalence documented from the Turks and Caicos (21 percent), Cayman 
Islands (20 percent), U.S. Virgin Islands (13 percent), and Cuba (8 
percent). No disease was recorded on A. cervicornis in Jamaica, Mexico, 
Netherlands Antilles, Panama, and Venezuela during the 1997 to 2000 
AGRRA surveys (Acropora Biological Review Team, 2005). Between 2001 and 
2002, disease was detected at all monitored thickets off Ft. 
Lauderdale, Florida with mortality ranging from 0.1 to 7.5 percent per 
site and a mean of 1.8 percent of colony surface area affected (Vargas-
Angel et al., 2003). Evidence of white band disease was observed on 5.3 
percent of A. cervicornis colonies in February 2010 at Cabezos del 
Cayo, Dominican Republic (Lirman et al., 2010a). During a disease 
outbreak in the Florida Keys in 2003, 72 percent of the 20 tagged A. 
cervicornis colonies were infected; 28 percent of these suffered 
complete mortality while many more colonies ended up as remnants of 
live tissue (less than 10 percent of colony alive; Williams and Miller, 
2005).
    The public comments provided the following supplemental information 
on the susceptibility of A. cervicornis to disease. No disease was 
detected in stratified random surveys in the Florida Keys in 2007 
(Miller et al., 2013). Vollmer and Kline (2008) found that six percent 
of A. cervicornis genotypes (three out of 49) were resistant to white 
band disease during in situ transmission assays in Bocas del Toro, 
Panama.
    Supplemental information we found on the susceptibility of A. 
cervicornis to disease includes the following. In Honduras, diseases 
were present in 32 percent of colonies (n = 181) monitored annually 
from 1996 to 2005 (Riegl et al., 2009). Between zero and 30 percent of 
A. cervicornis colonies monitored in the middle Florida Keys were 
affected by disease from 2011 to 2012 (Lunz, 2013). About five percent 
were affected by rapid tissue loss during each quarterly monitoring 
period (Lunz, 2013).
    All information on the susceptibility of A. cervicornis to disease 
can be summarized as follows. Acropora cervicornis is highly 
susceptible to disease as evidenced by the mass-mortality event in the 
1970s and 1980s. Although disease is both spatially and temporally 
variable, about five to six percent of A. cervicornis colonies appear 
to be affected by disease at any one time, though incidence of disease 
has been reported to range from zero to 32 percent and up to 72 percent 
during an outbreak. There is indication that some colonies may be 
resistant to white band disease. Acropora cervicornis is also 
susceptible to several diseases including one that causes rapid tissue 
loss from multi-focal lesions. Because few studies track diseased 
colonies over time, determining the present-day colony and population 
level effects of disease is difficult. One study that monitored 
individual colonies during an outbreak found that disease can be a 
major cause of both partial and total colony mortality (Williams and 
Miller, 2005). Thus, we conclude that A. cervicornis is highly 
susceptible to disease.
    Information on A. cervicornis' susceptibility to ocean warming that 
we considered in the proposed rule includes the following. Acropora 
cervicornis was one of the most heavily affected species during a 1987 
to 1988 bleaching event in the Cayman Islands with 100 percent of 
colonies bleached on the deep reef terrace (18 to 29 m depth) and 83 
percent bleached on the shallow reef terrace (Ghiold and Smith, 1990). 
In Puerto Rico, about 75 percent of A. cervicornis colonies bleached at 
12 monitored sites during the 2005 Caribbean bleaching event (Waddell 
and Clarke, 2008). At Culebra Island, Puerto Rico approximately 90 
percent of the A. cervicornis colonies had partial or total mortality 
during and after the 2005 bleaching event, and bleaching stress and 
mortality are believed to have resulted in the reproductive failure to 
subsequently spawn in 2006 (Waddell and Clarke, 2008).
    Repeat sampling of colonies in the Florida Keys and Bahamas in 
1998, and seasonally between March 2000 and August 2004, showed that 
colonies of A. cervicornis were stable with their associations with 
Symbiodinium type A3 but sometimes had mixed symbiosis with 
Symbiodinium type (B1) (Thornhill et al., 2006). The associations with 
Symbiodinium type (B1) were always short-lived (gone by next sampling 
period) and did not appear to be correlated with seasonal fluctuations 
or to follow the 1997 to 1998 bleaching event (Thornhill et al., 2006). 
Most of the mixed symbiosis events were limited to a single colony 
except for one sampling period in August 2001 when all colonies at one 
of the Bahamian sites had mixed symbionts.
    The public comments did not provide new or supplemental information 
on the susceptibility of A. cervicornis to ocean warming. Supplemental 
information we found on the susceptibility of A. cervicornis to ocean 
warming includes the following. In Roatan, Honduras, Riegl et al. 
(2009) monitored A. cervicornis and found none were bleached fully 
during the 1998 bleaching event, with the fourth highest partial 
bleaching frequency, and the highest mortality of 22 species monitored. 
During the 2005 bleaching event with 17 species observed, only A. 
cervicornis and A. palmata bleached 100 percent (all colonies bleached 
completely white) at two reefs in Jamaica with 90 percent mortality at 
one site and 10 percent at the other (Quinn and Kojis, 2008).
    Van Woesik et al. (2012) developed a coral resiliency index based 
on biological traits and processes to evaluate extinction risk due to 
bleaching. Evaluations were performed at the genus level with genera 
separated between the Caribbean and Indo-Pacific. They reported A. 
cervicornis as highly vulnerable to extinction due to bleaching.
    All information on the susceptibility of A. cervicornis to ocean 
warming can be summarized as follows. Acropora cervicornis is highly 
susceptible to bleaching in comparison to other coral species, and 
mortality after bleaching events is variable. Algal symbionts did not 
shift in A. cervicornis after the 1998 bleaching event, indicating the 
ability of this species to acclimatize to rising temperatures may not 
occur through this mechanism. Data from Puerto Rico and Jamaica 
following the 2005 Caribbean bleaching event indicate that temperature 
anomalies can have a large impact on total and partial mortality and 
reproductive output. Thus, we conclude that A. cervicornis is highly 
susceptible to ocean warming.

[[Page 53962]]

    Information on A. cervicornis' susceptibility to acidification that 
we considered in the proposed rule includes the following. Renegar and 
Riegl (2005) performed laboratory experiments to examine the effect of 
nutrients and carbon dioxide on A. cervicornis growth. They found 
significantly reduced growth under carbon dioxide levels of 700 to 800 
[mu]atm, predicted to occur this century, compared to controls. In 
addition, when elevated carbon dioxide was combined with increased 
nitrate and phosphate, growth rates were further reduced. The effect of 
combined nitrate, phosphate, and carbon dioxide appeared to be 
antagonistic at lower nutrient concentrations and additive at higher 
concentrations (compared to those nutrients paired with carbon dioxide 
separately). All corals in the combined nitrate, phosphate, and carbon 
dioxide treatment experienced total mortality, indicating the severe 
stress this combination induced.
    The public comments did not provide new or supplemental information 
on the susceptibility of A. cervicornis to acidification. Supplemental 
information we found on the susceptibility of A. cervicornis to 
acidification includes the following. Enochs et al. (2014) examined the 
effects of carbon dioxide and light intensity on A. cervicornis. They 
found that carbon dioxide levels projected to occur by the end of the 
century from ocean acidification caused reduced calcification and 
skeletal density but no change in linear extension, surface area, or 
volume. High light intensity did not ameliorate reductions in 
calcification, and the authors concluded that the high light intensity 
necessary to reach saturation of photosynthesis and calcification in A. 
cervicornis may limit the effectiveness of this potentially protective 
mechanism.
    All information on the susceptibility of A. cervicornis to 
acidification can be summarized as follows. Acropora cervicornis is 
susceptible to acidification through reduced growth, calcification, and 
skeletal density, and the effects of increased carbon dioxide combined 
with increased nutrients appear to be much worse than either stressor 
alone, and caused 100 percent mortality in some combination in one 
laboratory study. Therefore, we conclude that A. cervicornis is highly 
susceptible to acidification.
    There is no species-specific information on the trophic effects of 
fishing on A. cervicornis. However, due to the level of reef fishing 
conducted in the Caribbean, coupled with Diadema die-off and lack of 
significant recovery, recruitment habitat is limited. Thus, we conclude 
that A. cervicornis has some susceptibility to the trophic effects of 
fishing due to its low recruitment rates. However, the available 
information does not support a more precise description of 
susceptibility to this threat.
    All information on A. cervicornis' susceptibility to sedimentation 
that we considered in the proposed rule includes the following. 
Exposure to drilling mud reduced calcification rates and protein 
concentrations in A. cervicornis, and exposure to equivalent 
concentrations of kaolin produced no drop in proteins and a lower drop 
in calcification rate, indicating the toxic effects of drilling mud are 
not due solely to increases in turbidity (Kendall et al., 1983).
    Acropora cervicornis has poor capacity to remove coarser sediments 
(250-2000 [mu]m) and only slightly more capacity for removing finer 
sediments (62-250 [mu]m; Hubbard and Pocock, 1972). Water movement 
(turbulence) and gravity are probably more important in removing 
sediments from this species than its capabilities of sloughing 
sediments in still water (Porter, 1987). In field experiments in Puerto 
Rico, A. cervicornis was less sensitive to single applications (200 mg 
per cm\2\, 400 mg per cm\2\, and 800 mg per cm\2\) of coarse sediment 
(mean grain size 0.5 mm) than A. palmata and Orbicella annularis, 
likely due to morphology that facilitated passive sediment removal, 
though some bleaching near the base of the colonies did occur (Rogers, 
1983).
    Lab experiments testing the effects of sedimentation and phosphate 
on A. cervicornis indicated that sediment-clearing rates declined with 
increased exposure from less than two hours to up to 24 hours after 
four weeks of treatment. Treatments resulted in degenerative changes to 
tissue, zooxanthellae, and gonad development and were more severe in 
sediment and sediment plus phosphate treatments in comparison to 
controls and phosphate alone (Hodel and Vargas-Angel, 2007).
    Acropora cervicornis is sensitive to turbidity because it is highly 
reliant on sunlight for nutrition (Lewis, 1977; Porter, 1976). Rogers 
(1979) shaded a 20 m\2\ area of reef as a partial simulation of 
conditions from turbidity and found that A. cervicornis was the first 
species to respond by bleaching. Three weeks after shading was 
initiated, most colonies of A. cervicornis were bleached. After shading 
was terminated at five weeks, at the sixth week, most branches were 
dead and covered with algae with growth tips deteriorating or grazed 
away, but a few branches recovered. After seven weeks, there were more 
algae on the branches and further disintegration of branch tips.
    Fossil material collected from Bocas del Toro, Panama indicated 
that A. cervicornis declined in lagoonal areas prior to 1960, 
coincident with intensive land clearing, and continued to decline 
offshore after 1960, with community structure more tolerant of turbid 
conditions (Cramer et al., 2012).
    The public comments did not provide new or supplemental information 
on A. cervicornis' susceptibility to sedimentation, and we did not find 
any new or supplemental information. All information on the 
susceptibility of A. cervicornis to sedimentation can be summarized as 
follows. Acropora cervicornis is susceptible to sedimentation through 
its sensitivity to turbidity, and increased run-off from land clearing 
has resulted in mortality of this species. In addition, laboratory 
studies indicate the combination of sedimentation and nutrient 
enrichment appears to be worse than the effects of either of these two 
stressors alone. Thus, we conclude that A. cervicornis has high 
susceptibility to sedimentation.
    Information on A. cervicornis' susceptibility to nutrient 
enrichment that we considered in the proposed rule includes the 
following. Renegar and Riegl (2005) performed laboratory experiments to 
examine the effect of nutrients and carbon dioxide on A. cervicornis 
growth. Under the nutrient treatments alone, A. cervicornis experienced 
significantly lower growth rates under the higher nitrate and higher 
phosphate treatments, though not under the lower levels, and the 
combined nitrate and phosphate treatment produced significantly lower 
growth under both the low and high levels. All corals in the combined 
nitrate, phosphate, and carbon dioxide treatment experienced total 
mortality, indicating the severe stress this combination induced.
    Lab experiments testing the effects of sedimentation and phosphate 
on A. cervicornis indicated that degenerative changes to tissue, 
zooxanthellae, and gonad development were more severe in sediment plus 
phosphate treatments in comparison to controls and phosphate alone 
(Hodel and Vargas-Angel, 2007).
    The public comments did not provide new or supplemental information 
on the susceptibility of A. cervicornis to nutrient enrichment, and we 
did not find any new or supplemental information on its susceptibility 
to this threat. All information on the susceptibility of A. cervicornis 
to nutrient enrichment can be summarized as follows. Elevated nutrients 
can cause decreased growth in A. cervicornis. The

[[Page 53963]]

combined effects of nutrients with other stressors such as elevated 
carbon dioxide and sedimentation appear to be worse than the effects of 
nutrients alone, and can cause colony mortality in some combinations. 
Thus, we conclude that A. cervicornis is highly susceptible to nutrient 
enrichment.
    Information on A. cervicornis' susceptibility to predation that we 
considered in the proposed rule includes the following. Known predators 
of A. cervicornis include the corallivorous snail Coralliophila 
abbreviata and the polychaete fireworm Hermodice carunculata. Fireworms 
engulf growing branch tips and devour the live tissue; removal of 
tissue from growing branch tips of A. cervicornis may negatively affect 
colony growth. Corallivorous snails have also been shown to transmit a 
disease that causes rapid tissue loss in A. cervicornis (Williams and 
Miller, 2005). Several species of fish including, threespot damselfish 
Stegastes planifrons and yellowtail damselfish Microspathodon 
chrysurus, do not directly feed on coral but remove live tissue to 
cultivate algal gardens.
    In all thickets monitored off Ft. Lauderdale, Florida between 2001 
and 2002, densities of fireworms ranged between 18 and 86 individuals 
per hectare, with predation scars affecting less than 0.2 percent of 
the A. cervicornis cover (Vargas-Angel et al., 2003). Within the survey 
quadrats, fireworm scar sizes ranged from 1.0 to 8.0 cm, and densities 
ranged from 0 to 30 per m\2\ (Vargas-Angel et al., 2003). Evidence of 
fireworm predation was observed on 20.3 percent of colonies in Cabezos 
del Cayo, Dominican Republic in 2010 (Lirman et al., 2010a). Yellowtail 
damselfish and three-spot damselfish were present on A. cervicornis 
colonies at a density of 0.50 and 0.96 fish per m\2\, respectively, in 
the Dry Tortugas National Park, near Garden Key, Florida in 2004 
(Wilkes et al., 2008).
    The public comments provided the following supplemental information 
on the susceptibility of A. cervicornis to predation. In stratified 
random samples in the Florida Keys, damselfish algal gardens were 
detected on 1.9 percent of colonies in 2007 and 2.6 percent of colonies 
in 2012. Snail predation was detected on 1.3 percent of colonies in 
2012 (Miller et al., 2013).
    Supplemental information we found on the susceptibility of A. 
cervicornis to predation includes the following. In Cabezos del Cayo, 
Dominican Republic, 30 percent of colonies occurred within established 
damselfish territories, and corallivorous snails were found on 11.3 
percent of A. cervicornis colonies in 2010 (Lirman et al., 2010a). In 
permanent monitoring plots in the middle Florida Keys between 2011 and 
2012, about ten percent of fate-tracked A. cervicornis colonies were 
affected by fireworm predation, about five percent were affected by 
damselfish, and about five percent were affected by corallivorous 
snails (Lunz, 2013).
    All information on the susceptibility of A. cervicornis to 
predation can be summarized as follows. Predators can have a negative 
impact on A. cervicornis through both tissue removal and the spread of 
disease. Predation pressure appears spatially variable. Removal of 
tissue from growing branch tips of A. cervicornis may negatively affect 
colony growth, but the impact is unknown as most studies do not report 
on the same colonies through time, inhibiting evaluation of the longer-
term impact of these predators on individual colonies and populations. 
We conclude that A. cervicornis is highly susceptible to predation.
    Information on A. cervicornis' susceptibility to collection and 
trade that we considered in the proposed rule includes the following. 
Over the last decade, collection and trade of this species has been 
low.
    The public comments did not provide new or supplemental information 
on the susceptibility of A. cervicornis to collection and trade. 
Supplemental information we found includes the following. Over the last 
decade, collection and trade of this species has been primarily for 
scientific research rather than commercial purposes. Gross exports 
averaged approximately 2,500 pieces of coral per year between 2000 and 
2012 (data available at http://trade.cites.org). We conclude that A. 
cervicornis has low susceptibility to collection and trade.
    There is no species-specific information on the susceptibility of 
A. cervicornis to sea level rise. The SRR described sea level rise as 
an overall low to medium threat for all coral species. The public 
comments did not provide new or supplemental information on A. 
cervicornis' susceptibility to sea level rise, and we did not find any 
new or supplemental information. Thus, we conclude that A. cervicornis 
has some susceptibility to sea level rise, but the available 
information does not support a more precise description of 
susceptibility to this threat.
Regulatory Mechanisms
    In the proposed rule, we relied on information from the Final 
Management Report for evaluating the existing regulatory mechanisms for 
controlling threats to all corals. However, we did not provide any 
species-specific information on the regulatory mechanisms or 
conservation efforts for A. cervicornis. Public comments were critical 
of that approach, and we therefore attempt to analyze regulatory 
mechanisms and conservation efforts on a species basis, where possible, 
in this final rule. We also incorporate here, the evaluation of threats 
to this species conducted in the 2005 status review. Records confirm 
that A. cervicornis occurs in seven Atlantic ecoregions that encompass 
26 kingdom's and countries' EEZs. The 26 kingdoms and countries are 
Antigua & Barbuda, Bahamas, Barbados, Belize, Colombia, Costa Rica, 
Cuba, Dominica, Dominican Republic, French Antilles, Grenada, 
Guatemala, Haiti, Kingdom of the Netherlands, Honduras, Jamaica, 
Mexico, Nicaragua, Panama, St. Kitts & Nevis, St. Lucia, St. Vincent & 
Grenadines, Trinidad and Tobago, United Kingdom (British Caribbean 
Territories), United States (including U.S. Caribbean Territories), and 
Venezuela. The regulatory mechanisms relevant to A. cervicornis, 
described first as a percentage of the above kingdoms and countries 
that utilize them to any degree, and second as the percentages of those 
kingdoms and countries whose regulatory mechanisms may be limited in 
scope, are as follows: General coral protection (31 percent with 12 
percent limited in scope), coral collection (50 percent with 27 percent 
limited in scope), pollution control (31 percent with 15 percent 
limited in scope), fishing regulations on reefs (73 percent with 50 
percent limited in scope), managing areas for protection and 
conservation (88 percent with 31 percent limited in scope). The most 
common regulatory mechanisms in place for A. cervicornis are fishing 
regulations and area management for protection and conservation. 
However, half of the fishing regulations are limited in scope and may 
not provide substantial protection for the species. General coral 
protection and collection laws, along with pollution control laws, are 
much less common regulatory mechanisms for the management of A. 
cervicornis. The 2005 status review and 2006 listing concluded that 
existing regulatory mechanisms are inadequate to control both global 
and local threats, and are contributing to the threatened status of the 
species, and we incorporate that analysis here.
    Additionally, the public comments suggested that we did not fully 
consider the effects that conservation efforts have on the status of A. 
cervicornis. Therefore, conservation efforts are

[[Page 53964]]

described as follows. Conservation efforts have been underway for A. 
cervicornis for a number of years. Of 60 Acropora restoration efforts 
identified in 14 Caribbean countries, 88 percent used A. cervicornis 
including efforts in Belize, Colombia, Cura[ccedil]ao, Dominican 
Republic, Guadalupe, Honduras, Jamaica, Mexico, Puerto Rico, Turks and 
Caicos, U.S. Virgin Islands, and the Florida Keys (Young et al., 2012). 
The most popular method is to use coral nurseries to propagate A. 
cervicornis for restoration (Johnson et al., 2011; Young et al., 2012). 
Fast growth rates, branching morphology, and asexual reproduction 
through fragmentation make A. cervicornis an ideal candidate for active 
propagation. The use of coral nursery techniques has been shown to be 
effective and only temporarily affect wild donor colonies from which 
fragments are taken to initially stock nurseries (Lirman et al., 
2010b). Survivorship is high (greater than 70 percent) in nurseries 
during the first year, but mortality due to storms, temperature 
anomalies, predation, and water quality have been reported (Young et 
al., 2012). Survival rates are variable after transplanting, ranging 
between 43 and 95 percent during the first year (Hollarsmith et al., 
2012; Young et al., 2012). Mortality rates of non-nursery raised 
transplanted A. cervicornis after five years were similar to those of 
reference or wild colonies (Garrison and Ward, 2008).
    In conclusion, there are many conservation efforts aimed at 
increasing abundance and diversity of A. cervicornis throughout the 
Caribbean. These efforts are important, but not enough to ensure 
conservation unless combined with efforts to reduce the underlying 
threats and causes of mortality (Young et al., 2012). Thus, while 
conservation efforts will likely enhance recovery and conservation of 
A. cervicornis at small spatial scales, they are unlikely to affect the 
overall status of the species, given the global nature of threats.
Vulnerability to Extinction
    In 2006, A. cervicornis was listed as threatened, i.e., likely to 
become in danger of extinction within the next 30 years, due to: (1) 
Recent drastic declines in abundance of the species that have occurred 
throughout its geographic range and abundances at historic lows; (2) 
the potential constriction of broad geographic ranges due to local 
extirpations resulting from a single stochastic event (e.g., 
hurricanes, new disease outbreak); (3) limited sexual recruitment in 
some areas and unknown levels in most; and (4) occurrence of the Allee 
effect (in which fertilization success declines greatly as adult 
density declines).
    The species was not listed as endangered, i.e., currently in danger 
of extinction, because: (1) It was showing limited, localized recovery; 
(2) range-wide, the rate of decline appeared to have stabilized and was 
comparatively slow as evidenced by persistence at reduced abundances 
for the past two decades; (3) it was buffered against major threats by 
the large number of colonies, large geographic range, and asexual 
reproduction; and (4) as shown by the geologic record, the species has 
persisted through climate cooling and heating fluctuation periods over 
millions of years, whereas other corals have gone extinct.
    In 2012, A. cervicornis was proposed for listing as endangered 
because information available since the original 2006 listing as 
threatened suggested: (1) Population declines have continued to occur, 
with certain populations of both species decreasing up to an additional 
50 percent or more since the time of listing; (2) there are documented 
instances of recruitment failure in some populations; (3) minimal 
levels of thermal stress (e.g., 30 degrees C) have been shown to impair 
larval development, larval survivorship, and settlement success of A. 
palmata; (4) near-future levels of acidification have been demonstrated 
to impair fertilization, settlement success, and post-settlement growth 
rates in A. palmata; (5) on average 50 percent of the colonies are 
clones, meaning the effective number of genetic individuals is half the 
total population size; (6) the species' ranges are not known to have 
contracted, but with continued declines local extirpations are likely, 
resulting in a reduction of absolute range size. Furthermore, we took 
into account that the BRT identified restriction to the Caribbean as a 
spatial factor increasing extinction risk, though, among other things, 
exposure to high levels of human disturbance that result in pollution 
and breakage impacts. Also, while asexual reproduction (fragmentation) 
provides a source for new colonies (albeit clones) that can buffer 
natural demographic and environmental variability remains true, we 
believed that reliance on asexual reproduction is not sufficient to 
prevent extinction of the species. Last, the previous status review and 
listing determination underestimated the global climate change-
associated impacts to A. palmata and A. cervicornis, based on our 
current knowledge of trends in emissions, likely warming scenarios, and 
ocean acidification. In particular, in the previous determination, we 
identified ocean acidification only as a factor that ``may be 
contributing'' to the status of two species, in comparison to our 
current understanding that ocean acidification is one of the three 
highest order threats affecting extinction risk for corals.
    As explained above in the Risk Analyses section, a species' 
vulnerability to extinction results from the combination of its spatial 
and demographic traits, threat susceptibilities, and consideration of 
the baseline environment and future projections of threats. Subsequent 
to the proposed rule, we received and gathered supplemental species- or 
genus-specific information, described above, that expands our knowledge 
regarding the species' abundance, distribution, and threat 
susceptibilities. We developed our assessment of the species' 
vulnerability to extinction using all the available information. As 
explained in the Risk Analyses section, our assessment in this final 
rule emphasizes the ability of the species' spatial and demographic 
traits to moderate or exacerbate its vulnerability to extinction, as 
opposed to the approach we used in the proposed rule, which emphasized 
the species' susceptibility to threats.
    The following characteristics of A. cervicornis, in conjunction 
with the information described in the Corals and Coral Reefs section, 
Coral Habitat sub-section, and Threats Evaluation section above, affect 
its vulnerability to extinction currently and over the foreseeable 
future. The species has undergone substantial population decline and 
decreases in the extent of occurrence throughout its range due mostly 
to disease. Although localized mortality events have continued to 
occur, percent benthic cover and proportion of reefs where A. 
cervicornis is dominant have remained stable over its range since the 
mid-1980s. There is evidence of synergistic effects of threats for this 
species including worse effects of nutrients in combination with 
acidification and sedimentation. Acropora cervicornis is highly 
susceptible to a number of threats, and cumulative effects of multiple 
threats are likely to exacerbate vulnerability to extinction. Despite 
the large number of islands and environments that are included in the 
species' range, geographic distribution in the highly disturbed 
Caribbean exacerbates vulnerability to extinction over the foreseeable 
future because A. cervicornis is limited to an area with high, 
localized human impacts and predicted increasing threats. Acropora 
cervicornis

[[Page 53965]]

commonly occurs in water ranging from 5 to 20 m in depth, though occurs 
in deeper depths of 16-30 m at the northern extent of its range, and 
has been rarely found to 60 m in depth. It occurs in spur and groove, 
bank reef, patch reef, and transitional reef habitats, as well as on 
limestone ridges, terraces, and hardbottom habitats. This habitat 
heterogeneity moderates vulnerability to extinction over the 
foreseeable future because the species occurs in numerous types of reef 
environments that are predicted, on local and regional scales, to 
experience highly variable thermal regimes and ocean chemistry at any 
given point in time. Its absolute population abundance has been 
estimated as at least tens of millions of colonies in the Florida Keys 
and Dry Tortugas combined and is higher and is higher than the estimate 
from these two locations due to the occurrence of the species in many 
other areas throughout its range. Acropora cervicornis has low sexual 
recruitment rates, which exacerbates vulnerability to extinction due to 
decreased ability to recover from mortality events when all colonies at 
a site are extirpated. In contrast, its fast growth rates and 
propensity for formation of clones through asexual fragmentation 
enables it to expand between rare events of sexual recruitment and 
increases its potential for local recovery from mortality events, thus 
moderating vulnerability to extinction. Its abundance and life history 
characteristics, combined with spatial variability in ocean warming and 
acidification across the species' range, moderate vulnerability to 
extinction because the threats are non-uniform, and there will likely 
be a large number of colonies that are either not exposed or do not 
negatively respond to a threat at any given point in time.
Listing Determination
    In the proposed rule, using the determination tool formula 
approach, A. cervicornis was proposed for listing as endangered because 
of: High vulnerability to ocean warming (E), ocean acidification (E) 
and disease (C); high vulnerability to sedimentation (A and E) and 
nutrient over-enrichment (A and E); uncommon abundance (E); decreasing 
trend in abundance (E); low relative recruitment rate (E); narrow 
overall distribution (E); restriction to the Caribbean (E); and 
inadequacy of regulatory mechanisms (D).
    In this final rule, we changed the listing determination for A. 
cervicornis from endangered to threatened. We made this determination 
based on a more species-specific and holistic approach, including 
consideration of the buffering capacity of this species' spatial and 
demographic traits, and the best available information above on A. 
cervicornis' spatial structure, demography, threat susceptibilities, 
and management. The combination of these factors indicates that A. 
cervicornis is likely to become endangered throughout its range within 
the foreseeable future, and thus warrants listing as threatened at this 
time, because:
    (1) Acropora cervicornis is highly susceptible to ocean warming 
(ESA Factor E), disease (C), ocean acidification (E), sedimentation (A, 
E), nutrients (A, E), and predation (C) and susceptible to trophic 
effects of fishing (A), depensatory population effects from rapid, 
drastic declines and low sexual recruitment (E), and anthropogenic and 
natural abrasion and breakage (A, E). These threats are expected to 
continue and increase into the future. In addition, the species is at 
heightened extinction risk due to inadequate existing regulatory 
mechanisms to address both local and global threats (D);
    (2) Acropora cervicornis is geographically located in the highly 
disturbed Caribbean where localized human impacts are high and threats 
are predicted to increase as described in the Threats Evaluation 
section. A range constrained to this particular geographic area that is 
likely to experience severe and increasing threats indicates that a 
high proportion of the population of this species is likely to be 
exposed to those threats over the foreseeable future; and
    (3) Acropora cervicornis' abundance is still a fraction of what it 
was before the mass mortality in the 1970s and 1980s, and its presence 
on reefs throughout its range has continued to decrease over the last 
decade.
    The combination of these characteristics and future projections of 
threats indicates that the species is likely to be in danger of 
extinction within the foreseeable future throughout its range and 
warrants listing as threatened at this time due to factors A, C, D, and 
E.
    The available information above on A. cervicornis' spatial 
structure, demography, threat susceptibilities, and management also 
indicate that the species is not currently in danger of extinction and 
thus does not warrant listing as Endangered because:
    (1) While A. cervicornis' distribution within the Caribbean 
increases its risk of exposure to threats as described above, its 
habitat includes spur and groove, bank reef, patch reef, and 
transitional reef habitats, as well as limestone ridges, terraces, and 
hardbottom habitats in water depths ranging from 5 to 60 m. This 
moderates vulnerability to extinction currently because the species is 
not limited to one habitat type but occurs in numerous types of reef 
environments that will experience highly variable thermal regimes and 
ocean chemistry on local and regional scales at any given point in 
time, as described in more detail in the Coral Habitat and Threats 
Evaluation sections;
    (2) Acropora cervicornis' absolute abundance is at least tens of 
millions of colonies based on estimates from two locations. Absolute 
abundance is higher than estimates from these locations since A. 
cervicornis occurs in many other locations throughout its range, 
including a few small pockets of robust remnant populations. This 
absolute abundance allows for variation in the responses of individuals 
to threats to play a role in moderating vulnerability to extinction for 
the species to some degree, as described in more detail in the Corals 
and Coral Reefs section;
    (3) Recent information indicates that percent cover and proportions 
of Caribbean sites where A. cervicornis is dominant have stabilized;
    (4) Acropora cervicornis shows evidence of limited population 
expansion in some portions of its range under some circumstances (e.g., 
Dry Tortugas, southeast Florida); and
    (5) Acropora cervicornis has fast growth rates and high capacity to 
produce clones through asexual fragmentation, which can aid in recovery 
from mortality events.
    The combination of these characteristics indicates that the species 
does not exhibit the characteristics of one that is currently in danger 
of extinction, as described previously in the Risk Analyses section, 
and thus does not warrant listing as endangered at this time. 
Therefore, we withdraw our proposal to list A. cervicornis as 
endangered.
    Progress has been made with A. cervicornis-specific conservation 
and restoration projects, albeit small-scale, and these projects are 
likely to increase in the future. Within some countries, A. 
cervicornis-specific conservation and restoration projects show promise 
for enhancing species recovery at very small spatial scales and for 
facilitating the persistence of the species in some areas in the face 
of continuing threats. Range-wide, a multitude of conservation efforts 
are already broadly employed specifically for A. cervicornis. However, 
considering the global scale of the most important threats to the 
species, and the ineffectiveness of conservation efforts at addressing 
the root cause of global threats (i.e., GHG emissions), we do not 
believe that any current conservation

[[Page 53966]]

efforts or conservation efforts planned in the future will result in 
affecting the species' status to the point at which listing is not 
warranted.

A. palmata

Introduction
    Acropora palmata colonies have frond-like branches, which appear 
flattened to near round, and typically radiate out from a central trunk 
and angle upward. Branches are up to 50 cm wide and range in thickness 
from 4 to 5 cm. Individual colonies can grow to at least 2 m in height 
and 4 m in diameter (Acropora Biological Review Team, 2005). Colonies 
of A. palmata can grow in nearly mono-specific, dense stands and form 
an interlocking framework known as thickets.
Spatial Information
    Information on A. palmata's distribution, habitat, and depth range 
that we considered in the proposed rule includes the following. 
Acropora palmata is distributed throughout the western Atlantic, 
Caribbean, and Gulf of Mexico. The northern extent of the range in the 
Atlantic is Broward County, Florida where it is relatively rare (only a 
few known colonies), but fossil A. palmata reef framework extends into 
Palm Beach County, Florida. There are two known colonies of A. palmata, 
which were discovered only recently in 2003 and 2005, at the Flower 
Garden Banks, located 161 km off the coast of Texas in the Gulf of 
Mexico (Zimmer et al., 2006).
    Acropora palmata often grows in thickets in fringing and barrier 
reefs (Jaap, 1984; Tomascik and Sander, 1987; Wheaton and Jaap, 1988) 
and formed extensive barrier-reef structures in Belize (Cairns, 1982), 
the greater and lesser Corn Islands, Nicaragua (Lighty et al., 1982), 
and Roatan, Honduras, and built extensive fringing reef structures 
throughout much of the Caribbean (Adey, 1978). Acropora palmata 
commonly grows in turbulent water on the fore-reef, reef crest, and 
shallow spur-and-groove zone (Cairns, 1982; Miller et al., 2008; Rogers 
et al., 1982; Shinn, 1963) in water ranging from 1 to 5 m depth. Early 
studies termed the reef crest and adjacent seaward areas from the 
surface to five or six meters depth the ``palmata zone'' because of the 
domination by the species (Goreau, 1959; Shinn, 1963). Maximum depth of 
framework construction ranges from 3 to 12 m, and colonies generally do 
not form thickets below a depth of 5 m (Lighty et al., 1982). Although 
A. palmata's predominant habitat is reef crests and shallow fore-reefs 
less than 12 m depth, it also occurs in back-reef environments and in 
depths up to 30 m.
    Extensive stands of dead colonies throughout the range occurred 
after mass mortalities during the 1970s and 1980s (see Demographic 
Information Below). There is no evidence of overall range constriction 
from the mass mortalities, but local extirpations are likely (Jackson 
et al., 2014), resulting in a reduction of absolute range size.
    The public comments did not provide new or supplemental information 
on A. palmata's habitat or depth range but provided the following 
supplemental information on its distribution. Precht and Aronson (2004) 
suggested that the recent expansion of A. palmata to the Flower Garden 
Banks (Zimmer et al., 2006) is possibly due to climate warming.
    Supplemental information we found on A. palmata's distribution is 
consistent with prior information. Veron (2014) confirms the occurrence 
of A. palmata in eight of a potential 11 ecoregions in the western 
Atlantic and wider-Caribbean that are known to contain corals. The 
three ecoregions in which A. palmata is not found are off the coasts of 
Bermuda, Brazil, and the southeast U.S. north of south Florida. The 
presence of the species in the Flower Garden Banks may represent a 
recent re-occupation of its historic range since fossil evidence 
indicates this species occupied the Flower Garden Banks during the 
early Holocene but disappeared in the middle Holocene due to sea level 
rise and possibly cooling temperatures (Precht et al., 2014). Finally, 
the spatial structure of the species has been affected by extirpation 
from many localized areas throughout its range (Jackson et al., 2014).
    Supplemental information we found on A. palmata's habitat and depth 
includes the following. Goreau (1959) described ten habitat zones on a 
Jamaican fringing reef from inshore to the deep slope, finding A. 
palmata in eight of the ten zones. Acropora palmata was very abundant 
in the reef crest zones, but also common in several other zones further 
inshore (the reef flat, rear, channel or lagoon, and inshore zones), 
and rare on the reef slope to 15 meters depth. Although A. palmata is 
currently much less common throughout its range than it was prior to 
the mid-1980s, it still occurs in multiple habitats and to depths of 
one to 30 m. For example, a 2005 study of Bonaire back-reefs found A. 
palmata at three of six sites, including within inshore and lagoon 
habitats, ranging from seven to 15 m depth. In 2003, aggregations of A. 
palmata were reported from patch reefs at 10 to 20 m depth within the 
lagoon of Serrano Bank (Sanchez and Pizarro, 2005).
Demographic Information
    Information on A. palmata's abundance and population trends that we 
considered in the proposed rule includes the following. Acropora 
palmata has been described as usually common (Veron, 2000) and uncommon 
(Carpenter et al., 2008). Acropora palmata was historically one of the 
dominant species on Caribbean reefs, forming large, monotypic thickets 
and giving rise to the nominal distinct zone in classical descriptions 
of Caribbean reef morphology (Goreau, 1959). Mass mortality, apparently 
from white-band disease (Aronson and Precht, 2001), spread throughout 
the Caribbean in the mid-1970s to mid-1980s and precipitated widespread 
and radical changes in reef community structure (Brainard et al., 
2011). This mass mortality occurred throughout the range of the species 
within all Caribbean countries and archipelagos, even on reefs and 
banks far from localized human influence (Aronson and Precht, 2001; 
Wilkinson, 2008). In addition, continuing coral mortality from periodic 
acute events such as hurricanes, disease outbreaks, and mass bleaching 
events added to the decline of A. palmata (Brainard et al., 2011). In 
locations where historic quantitative data are available (Florida, 
Jamaica, U.S. Virgin Islands), there was a reduction of greater than 97 
percent between the 1970s and early 2000s (Acropora Biological Review 
Team, 2005).
    Since the 2006 listing of A. palmata as threatened, continued 
population declines have occurred in some locations with certain 
populations of A. palmata and A. cervicornis decreasing up to an 
additional 50 percent or more (Colella et al., 2012; Lundgren and 
Hillis-Starr, 2008; Muller et al., 2008; Rogers and Muller, 2012; 
Williams et al., 2008). In addition, Williams et al. (2008) reported 
recruitment failure between 2004 and 2007 in the upper Florida Keys 
after a major hurricane season in 2005; less than five percent of the 
fragments produced recruited into the population.
    The public comments provided the following supplemental information 
on A. palmata's abundance and population trends. Several studies 
describe A. palmata populations that are showing some signs of recovery 
or are in good condition including in the Turks and Caicos Islands 
(Schelten et al., 2006), U.S. Virgin Islands (Grober-Dunsmore et al., 
2006; Mayor et al., 2006; Rogers and Muller, 2012), Venezuela 
(Zubillaga et

[[Page 53967]]

al., 2008), and Belize (Macintyre and Toscano, 2007).
    Extrapolated population estimates of A. palmata from stratified 
random samples across habitat types in the Florida Keys were 0.6  0.5 million (SE) colonies in 2005, 1.0  0.3 million 
(SE) colonies in 2007, and 0.5  0.3 million colonies in 
2012. Because these population estimates are based on random sampling, 
differences between years may be a function of sampling effort rather 
than an indication of population trends. Relative to the abundance of 
other corals in the Florida Keys region, A. palmata was among the least 
abundant, ranking among corals that are naturally rare in abundance. No 
colonies of A. palmata were observed in surveys of the Dry Tortugas in 
2006 and 2008. The size class distribution of the Florida Keys 
population included both small and large individuals (> 260 cm), but 
after 2005 the majority of the colonies were smaller in size. These 
smallest corals (0 to 20 cm) had approximately zero to two percent 
partial mortality during all three survey years. Partial mortality 
across all other size classes was approximately 20 to 70 percent in 
2005, 5 to 50 percent in 2007, and 15 to 90 percent in 2012 (Miller et 
al., 2013).
    Supplemental information we found on A. palmata's abundance 
includes the following. Relatively abundant A. palmata communities have 
been documented from various locations, including Cuba (Alcolado et 
al., 2010; Gonz[aacute]lez-D[iacute]az et al., 2010), Colombia (Sanchez 
and Pizarro, 2005), Venezuela (Mart[iacute]nez and Rodr[iacute]guez 
Quintal, 2012), Navassa (Bruckner, 2012b), Jamaica (Jackson et al., 
2014), and the U.S. Virgin Islands (Muller et al., 2014). Density 
estimates from sites in Cuba range from 0.14 colonies per m\2\ 
(Alcolado et al., 2010) to 0.18 colonies per m\2\ (Gonz[aacute]lez-
D[iacute]az et al., 2010). Maximum A. palmata density at ten sites in 
St. John, U.S. Virgin Islands was 0.18 colonies per m\2\ (Muller et 
al., 2014).
    Mayor et al. (2006) reported the abundance of A. palmata in Buck 
Island Reef National Monument, St. Croix, U.S. Virgin Islands. They 
surveyed 617 sites from May to June 2004 and extrapolated density 
observed per habitat type to total available habitat. Within an area of 
795 ha, they estimated 97,232-134,371 (95% confidence limits) A. 
palmata colonies with any dimension of connected live tissue greater 
than one meter. Mean densities (colonies >= 1 m) were 0.019 colonies 
per m\2\ in branching coral-dominated habitats and 0.013 colonies per 
m\2\ in other hard bottom habitats.
    Puerto Rico contains the greatest known extent of A. palmata in the 
U.S. Caribbean. Between 2006 and 2007, a survey of 431 random points in 
habitat suitable for A. palmata in six marine protected areas in Puerto 
Rico revealed a variable density of zero to 52 A. palmata colonies per 
100 m\2\ (0.52 colonies per m\2\), with average density of 3.3 colonies 
per 100 m\2\ (0.03 colonies per m\2\). Total loss of A. palmata was 
evidenced in 13.6 percent of the random survey areas where only dead 
standing colonies were present (Sch[auml]rer et al., 2009).
    In stratified random surveys along the south, southeast, southwest, 
and west coasts of Puerto Rico designed to locate Acropora colonies, A. 
palmata was observed at five out of 301 stations with sightings outside 
of the survey area at an additional two stations (Garc[iacute]a Sais et 
al., 2013). Acropora palmata colonies were absent from survey sites 
along the southeast coast. Maximum density was 18 colonies per 15 m\2\ 
(1.2 colonies per m\2\), and maximum colony size was 2.3 m in diameter 
(Garc[iacute]a Sais et al., 2013).
    Zubillaga et al. (2005) report densities of 3.2 colonies of A. 
palmata per 10 m\2\ (0.32 colonies per m\2\) in Los Roques National 
Park, Venezuela. At ten sites surveyed in the national park in 2003 to 
2004, density ranged from 0 to 3.4 colonies per 10 m\2\ (0 to 0.34 
colonies per m\2\) with four of the sites showing only standing dead 
colonies (Zubillaga et al., 2008). In the six sites with live colonies, 
small (0.1 to 50 cm\2\) and medium-sized (50 to 4,550 cm\2\) colonies 
predominated over larger-sized (4,550 to16,500 cm\2\) colonies.
    At Los Colorados reef in northwestern Cuba, a 2006 study at 12 reef 
crest sampling stations reported average A. palmata densities of 0.18 
colonies per m\2\, and that A. palmata made up 8.7 percent of the total 
live coral colonies at the study sites. The study also reported that 
the nearby Baracoa and Rincon de Guanabo reefs had similar A. palmata 
densities (Gonz[aacute]lez-D[iacute]az et al., 2010). The size of A. 
palmata colonies indicates some recruitment in Cuba, but not the 
proportions of sexual versus asexual recruits. In a 2005 study of 280 
A. palmata colonies at four sites on the north coast of Cuba, 30.4 
percent were less than 10 cm in diameter (Gonz[aacute]lez-D[iacute]az 
et al., 2008). In a 2006 study of approximately 1,100 A. palmata 
colonies at three sites on the north coast of Cuba, diameter and height 
size-classes were measured (<2, 3-5, 6-7, 8-10, 11-80, and >80 cm). For 
the three sites combined, there were approximately 25 to 100 colonies 
in each of the four smaller size classes (Perera-P[eacute]rez et al., 
2012).
    Supplemental information we found on A. palmata's population trends 
includes the following. At eight of 11 sites in St. John, U.S. Virgin 
Islands, colonies of A. palmata increased in abundance, between 2001 
and 2003, particularly in the smallest size class, with the number of 
colonies in the largest size class decreasing (Grober-Dunsmore et al., 
2006). Colonies of A. palmata monitored monthly between 2003 and 2009 
in Haulover Bay on St. John, U.S. Virgin Islands suffered bleaching and 
mortality from disease but showed an increase in abundance and size at 
the end of the monitoring period (Rogers and Muller, 2012). The overall 
density of A. palmata colonies around St. John did not significantly 
differ between 2004 and 2010 with six out of the ten sites showing an 
increase in colony density. Size frequency distribution did not 
significantly change at seven of the 10 sites, with two sites showing 
an increased abundance of large-sized (> 51 cm) colonies (Muller et 
al., 2014).
    In Colombia, A. palmata was present at four of the 32 plots (three 
of the six reefs) monitored annually from 1998 to 2004. Coverage of A. 
palmata ranged from 0.8 to 2.4 percent. Over the eight-year period, the 
species was stable at two reefs and declined at the other reef, likely 
in response to a hurricane in 1999 (Rodriguez-Ramirez et al., 2010). 
MacIntyre and Toscano (2007) report the return of ``numerous large 
colonies'' of A. palmata on the shallow fore-reef at the southern limit 
of Carrie Bow Cay, Belize though no quantitative data were presented.
    Colonies monitored in the upper Florida Keys showed a greater than 
50 percent loss of tissue as well as a decline in the number of 
colonies, and a decline in the dominance by large colonies between 2004 
and 2010 (Vardi et al., 2012; Williams and Miller, 2012). Elasticity 
analysis from a population model based on data from the Florida Keys 
has shown that the largest individuals have the greatest contribution 
to the rate of change in population size (Vardi et al., 2012). Between 
2010 and 2013 A. palmata in the middle and lower Florida Keys had mixed 
trends. Population densities remained relatively stable at two sites 
and decreased at two sites by 21 and 28 percent (Lunz, 2013).
    Acropora palmata monitored in Cura[ccedil]ao between 2009 and 2011 
decreased in abundance, increased in colony size, with stable tissue 
abundance following hurricane damage (Bright et al., 2013). The authors 
explained that the apparently conflicting trends of increasing colony

[[Page 53968]]

size but similar tissue abundance likely resulted from the loss of 
small-sized colonies that skewed the distribution to larger size 
classes, rather than colony growth.
    Simulation models using data from matrix models of A. palmata 
colonies from specific sites in Cura[ccedil]ao (2006-2011), the Florida 
Keys (2004-2011), Jamaica (2007-2010), Navassa (2006 and 2009), Puerto 
Rico (2007 and 2010), and the British Virgin Islands (2006 and 2007) 
indicate that most of these studied populations will continue to 
decline in size and extent by 2100 if background environmental 
conditions remain unchanged (Vardi, 2011). In contrast, the studied 
populations in Jamaica were projected to increase in abundance, and 
studied populations in Navassa were projected to remain stable. Studied 
populations in the British Virgin Islands were predicted to decrease 
slightly from their initial very low levels. Studied populations in 
Florida, Cura[ccedil]ao, and Puerto Rico were predicted to decline to 
zero by 2100. Because the study period did not include physical damage 
(storms), the population simulations in Jamaica, Navassa, and the 
British Virgin Islands may have contributed to the differing projected 
trends at sites in these locations.
    New information we found on population trends includes the 
following. A report on the status and trends of Caribbean corals over 
the last century indicates that cover of A. palmata has remained 
relatively stable at approximately one percent throughout the region 
since the large mortality events of the 1970s and 1980s. The report 
also indicates that the number of reefs with A. palmata present 
steadily declined from the 1980s to 2000-2004, then remained stable 
between 2000-2004 and 2005-2011. Acropora palmata was present at about 
20 percent of reefs surveyed in both the 5-year period of 2000-2004 and 
the 7-year period of 2005-2011. Acropora palmata was dominant on 
approximately five to ten percent of hundreds of reef sites surveyed 
throughout the Caribbean during the four periods of 1990-1994, 1995-
1999, 2000-2004, and 2005-2011 (Jackson et al., 2014).
    All information on A. palmata's abundance and population trends is 
summarized as follows. Based on population estimates there are at least 
hundreds of thousands of A. palmata colonies present in both the 
Florida Keys and St. Croix, U.S. Virgin Islands. Absolute abundance is 
higher than estimates from these two locations given the presence of 
this species in many other locations throughout its range. The 
effective population size is smaller than indicated by abundance 
estimates due to the tendency for asexual reproduction. Across the 
Caribbean, percent cover appears to have remained relatively stable 
since the population crash in the 1980s. Frequency of occurrence has 
decreased since the 1980s, indicating potential decreases in the extent 
of occurrence and effects on the species' range. However, the 
proportions of Caribbean sites where A. palmata is present and dominant 
have recently stabilized. There are locations such as the U.S. Virgin 
Islands where populations of A. palmata appear stable or possibly 
increasing in abundance and some such as the Florida Keys where 
population number appears to be decreasing. In some cases when size 
class distribution is not reported, there is uncertainty of whether 
increases in abundance indicate growing populations or fragmentation of 
larger size classes into more small-sized colonies. From locations 
where size class distribution is reported, there is evidence of 
recruitment, but not the proportions of sexual versus asexual recruits. 
The best evidence of recovery would come from multi-year studies 
showing an increase in the overall amount of living tissue of this 
species, growth of existing colonies, and an increase in the number of 
small corals arising from sexual recruitment (Rogers and Muller, 2012). 
Simulation models predict by 2100 that A. palmata will become absent at 
specific sites in several locations (Florida, Curacao, and Puerto 
Rico), decrease at specific sites in the British Virgin Islands, remain 
stable at specific sites in Navassa, and increase at specific sites in 
Jamaica. These simulations are based on the assumption that conditions 
experienced during the monitoring period, ranging from one to seven 
years depending on location, would remain unchanged in the future. We 
conclude there has been a significant decline of A. palmata throughout 
its range, with recent population stability at low percent coverage. We 
also conclude that absolute abundance is at least hundreds of thousands 
of colonies, but likely to decrease in the future with increasing 
threats.
Other Biological Information
    Information on A. palmata's life history that we considered in the 
proposed rule includes the following. Growth rates, measured as 
skeletal extension of the end of branches, range from 4 to 11 cm per 
year (Acropora Biological Review Team, 2005) but in Cura[ccedil]ao have 
been reported to be slower today than they were several decades ago 
(Brainard et al., 2011).
    Acropora palmata is a hermaphroditic broadcast spawning species 
that reproduces after the full moon of July, August, and/or September 
(Acropora Biological Review Team, 2005). The estimated size at sexual 
maturity is 1600 cm\2\, and growing edges and encrusting base areas are 
not fertile (Soong and Lang, 1992). Larger colonies have higher 
fecundity per unit area, as do the upper branch surfaces (Soong and 
Lang, 1992). Although self-fertilization is possible, A. palmata is 
largely self-incompatible (Baums et al., 2005a; Fogarty et al., 2012b).
    Reproduction occurs primarily through asexual fragmentation that 
produces multiple colonies that are genetically identical (Bak and 
Criens, 1982; Highsmith, 1982; Lirman, 2000; Miller et al., 2007; 
Wallace, 1985). Storms can be an important mechanism to produce 
fragments to establish new colonies (Fong and Lirman, 1995). 
Fragmentation is an important mode of reproduction in many reef-
building corals, especially for branching species such as A. palmata 
(Highsmith, 1982; Lirman, 2000; Wallace, 1985). However, in the Florida 
Keys where populations have declined, there have been reports of 
failure of asexual recruitment due to high fragment mortality after 
storms (Porter et al., 2012; Williams and Miller, 2010; Williams et 
al., 2008).
    Sexual recruitment rates are low, and this species is generally not 
observed in coral settlement studies. Laboratory studies have found 
that certain species of crustose-coralline algae facilitate larval 
settlement and post-settlement survival (Ritson-Williams et al., 2010). 
Rates of post-settlement mortality after nine months are high based on 
settlement experiments (Szmant and Miller, 2005).
    The public comments did not provide new or supplemental information 
on A. palmata's life history. Supplemental information we found on A. 
palmata's life history includes the following. Split spawning (spawning 
over a two month period) has been reported from the Florida Keys 
(Fogarty et al., 2012b). Laboratory experiments have shown that some 
individuals (i.e., genotypes) are sexually incompatible (Baums et al., 
2013) and that the proportion of eggs fertilized increases with higher 
sperm concentration (Fogarty et al., 2012b). Experiments using gametes 
collected in Florida had lower fertilization rates than those from 
Belize, possibly due to genotype incompatibilities (Fogarty et al., 
2012b).
    Darling et al. (2012) performed a biological trait-based analysis 
to

[[Page 53969]]

categorize coral species into four life history strategies: Generalist, 
weedy, competitive, and stress-tolerant. The classifications were 
primarily separated by colony morphology, growth rate, and reproductive 
mode. Acropora palmata was classified as a ``competitive'' species, 
thus likely more vulnerable to environmental stress.
    All information on A. palmata's life history can be summarized as 
follows. The combination of rapid skeletal growth rates and frequent 
asexual reproduction by fragmentation can enable effective competition 
within, and domination of, reef-building coral communities in high-
energy environments such as reef crests. Rapid skeletal growth rates 
and frequent asexual reproduction by fragmentation facilitate potential 
recovery from disturbances when environmental conditions permit 
(Highsmith, 1982; Lirman, 2000). However, low sexual reproduction can 
lead to reduced genetic diversity and limits the capacity to repopulate 
sites.
    Other biological information on A. palmata that we considered in 
the proposed rule includes the following. Genetic samples from 11 
locations throughout the Caribbean indicate that A. palmata populations 
in the eastern Caribbean (St. Vincent and the Grenadines, U.S. Virgin 
Islands, Cura[ccedil]ao, and Bonaire) have had little or no genetic 
exchange with populations in the western Atlantic and western Caribbean 
(Bahamas, Florida, Mexico, Panama, Navassa, and Puerto Rico) (Baums et 
al., 2005b). While Puerto Rico is more closely connected with the 
western Caribbean, it is an area of mixing with contributions from both 
regions (Baums et al., 2005b). Models suggest that the Mona Passage 
between the Dominican Republic and Puerto Rico acts as a filter for 
larval dispersal and gene flow between the eastern Caribbean and 
western Caribbean (Baums et al., 2006b).
    The western Caribbean is characterized by genetically depauperate 
populations with lower densities (0.13  0.08 colonies per 
m\2\), while denser (0.30  0.21 colonies per m\2\), 
genotypically rich stands characterize the eastern Caribbean (Baums et 
al., 2006a). Baums et al. (2006a) concluded that the western Caribbean 
had higher rates of asexual recruitment and that the eastern Caribbean 
had higher rates of sexual recruitment. They postulated these 
geographic differences in the contribution of reproductive modes to 
population structure may be related to habitat characteristics, 
possibly the amount of shelf area available.
    Genotypic diversity is highly variable. At two sites in the Florida 
Keys, only one genotype per site was detected out of 20 colonies 
sampled at each site (Baums et al., 2005b). In contrast, all 15 
colonies sampled in Navassa had unique genotypes (Baums et al., 2006a). 
Some sites have relatively high genotypic diversity such as in Los 
Roques, Venezuela (118 unique genotpyes out of 120 samples; Zubillaga 
et al., 2008) and in Bonaire and Curacao (18 genotypes of 22 samples 
and 19 genotypes of 20 samples, respectively; Baums et al., 2006a). In 
the Bahamas, about one third of the sampled colonies were unique 
genotypes, and in Panama between 24 and 65 percent of the sampled 
colonies had unique genotypes, depending on the site (Baums et al., 
2006a).
    The public comments did not provide new or supplemental biological 
information on A. palmata. Supplemental biological information we found 
includes the following. A genetic study found significant population 
structure in Puerto Rico locations (Mona Island, Desecheo Island, La 
Parguerain, La Parguera) both between reefs and between locations; 
population structure in La Parguera suggests restriction of gene flow 
between some reefs in close proximity (Garcia Reyes and Schizas, 2010). 
A more-recent study provided additional detail on the genetic structure 
of A. palmata in Puerto Rico, as compared to Curacao, the Bahamas, and 
Guadeloupe that found unique genotypes in 75 percent of the samples 
with high genetic diversity (M[egrave]ge et al., 2014). The recent 
results support two separate populations of A. palmata in the eastern 
Caribbean and western Caribbean; however, there is less evidence for 
separation at Mona Passage, as found by Baums et al. (2006b).
    All biological information on A. palmata can be summarized as 
follows. Genotypic diversity is variable across the range with some 
populations showing evidence of higher input from sexual recruitment 
versus others that rely more heavily on asexual recruitment for 
population maintenance. There are many areas with many unique 
genotypes. Connectivity and mixing appear limited across larger 
geographic scales with eastern Caribbean populations relatively 
isolated from western Caribbean populations, with evidence of 
population structure at a local scale in some locations.
Susceptibility to Threats
    Information on threat susceptibilities was interpreted in the 
proposed rule for A. palmata's vulnerability to threats as follows: 
High vulnerability to ocean warming, disease, acidification, 
sedimentation, and nutrient enrichment; moderate vulnerability to the 
trophic effects of fishing and predation; and low vulnerability to sea 
level rise and collection and trade.
    Information on A. palmata's susceptibility to disease that we 
considered in the proposed rule includes the following. Disease is 
believed to be the primary cause of the region-wide decline of A. 
palmata beginning in the late 1970s and continues to have a large 
effect on the species. White band disease was generally associated with 
the majority of disease-related mortalities in A. palmata from the 
1970s to 1990s (Aronson and Precht, 2001). White pox has been described 
as having severe impacts on A. palmata, and most monitoring information 
after 2000 indicates that lesion patterns resembling white pox have 
higher prevalence than patterns resembling white band disease (Acropora 
Biological Review Team, 2005). In the Florida Keys, the causative agent 
of white pox was identified as a bacterium linked to human sewage and 
potential vectors/reservoirs such as corallivores (Patterson et al., 
2002; Sutherland et al., 2011).
    The effects of white pox appear to be exacerbated by higher 
temperatures. In Hawksnest Bay, U.S. Virgin Islands during the 2005 
bleaching event, the prevalence of white pox had a positive linear 
relationship with temperature, with mortality increasing with 
bleaching, indicating a decreased resilience to disease when colonies 
were stressed (Muller et al., 2008).
    Disease is temporally and spatially variable and is often reported 
as an instantaneous measure of prevalence (percent of colonies affected 
by disease) that provides only a snapshot in time. For instance, in 
Puerto Rico disease affected an average of 6.7 percent of colonies from 
December 2006 to October 2007 (Sch[auml]rer et al., 2009). In St. Croix 
U.S. Virgin Islands, white band disease affected three percent of the 
colonies surveyed in Buck Island Reef National Monument between May and 
June 2004 (Mayor et al., 2006).
    Studies of permanently marked colonies, or monitoring plots, show 
longer-term trends of disease and mortality over time. From January 
2003 to December 2009, 90 percent of the 69 monitored A. palmata 
colonies in Haulover Bay, St. John, U.S. Virgin Islands exhibited signs 
of disease, and the most significant cause of whole colony mortality 
(Rogers and Muller, 2012). Of colonies monitored in the

[[Page 53970]]

Florida Keys from 2004 to 2011, disease was the second highest cause of 
tissue mortality after physical damage from storms (33 percent of all 
mortality attributed to disease, Williams and Miller, 2012).
    The public comments did not provide new or supplemental information 
on the susceptibility of A. palmata to disease, and we did not find any 
new or supplemental information. Information on the susceptibility of 
A. palmata to disease can be summarized as follows. Acropora palmata is 
highly susceptible to disease as evidenced by the mass-mortality event 
in the 1970s and 1980s. White pox seems to be more common today than 
white band disease. The effects of disease are spatially and temporally 
(both seasonally and inter-annually) variable. Results from longer-term 
monitoring studies in the U.S. Virgin Islands and the Florida Keys 
indicate that disease can be a major cause of both partial and total 
colony mortality. Thus, we conclude that A. palmata is highly 
susceptible to disease.
    Information on A. palmata's susceptibility to ocean warming that we 
considered in the proposed rule includes the following. High 
temperatures can cause bleaching and mortality of A. palmata. In St. 
Croix, U.S. Virgin Islands, colonies differentially bleached in Buck 
Island National Monument during the 2005 Caribbean-wide mass bleaching 
event; colonies in the shallower back reef bleached earlier and 
suffered greater tissue loss than those located elsewhere (Lundgren and 
Hillis-Starr, 2008). Data from two sites in Jamaica, found 100 percent 
of A. palmata colonies bleached at both sites in 2005, with greater 
than 50 percent of the colonies suffering partial mortality (Quinn and 
Kojis, 2008). At one site, bleached colonies had complete mortality 
only occasionally, and 15 percent of bleached colonies died at the 
second site (Quinn and Kojis, 2008). In Trunk Bay and Saltpond, St. 
John, U.S. Virgin Islands, almost half of the colonies that bleached in 
2005 suffered partial or complete mortality (44 percent of 27 colonies 
and 40 percent of 107 colonies, respectively, Rogers et al., 2006). 
Negligible bleaching of A. palmata was observed during a 2006 bleaching 
event in Navassa that affected corals at deeper depths (between 18 and 
37 m) more significantly than at shallower depths (<10 m), likely due 
to decreased water motion at the deeper sites (Miller et al., 2011a). 
Repeated sampling of the same colonies in the Florida Keys and Bahamas 
in 1998 and seasonally between March 2000 and August 2004 showed that 
colonies of A. palmata did not change their association with 
Symbiodinium type A3 throughout the study period that included the 
1997-98 bleaching event (Thornhill et al., 2006).
    High water temperatures also affect A. palmata reproduction. 
Acropora palmata embryos and larvae exhibited more developmental 
abnormalities, lower survivorship, and decreased settlement at 30 
degrees and 31.5 degrees C compared to those at 28 degrees C (Randall 
and Szmant, 2009).
    The public comments did not provide new or supplemental information 
on the susceptibility of A. palmata to ocean warming. Supplemental 
information we found includes the following. Acropora palmata larvae 
exhibited faster development and faster swimming speed at 30 and 31.5 
degrees C compared to controls at 27 and 28 degrees C (Baums et al., 
2013). The authors suggested these changes could decrease average 
larval dispersal distances, and cause earlier larval settlement, 
thereby affecting gene flow among populations (Baums et al., 2013).
    A 14-year study was conducted at nine sites around Little Cayman 
from 1999 to 2012 of live coral cover, juvenile densities, and size 
structure of coral colonies to determine response to the 1998 bleaching 
event inside versus outside of marine protected areas. Over the first 
half of the study, bleaching and disease caused live cover to decrease 
from 26 percent to 14 percent in all corals, with full recovery seven 
years later with no differences inside versus outside of marine 
protected areas. The numbers of A. palmata colonies in regularly-
observed size-classes did not decrease during this study, which the 
authors suggested may indicate resistance to bleaching and disease. The 
study concluded that the health of the coral assemblage and the 
similarity of responses inside and outside the marine protected area 
suggested that negligible anthropogenic disturbance at the local scale 
was a key factor underlying the observed resilience (Manfrino et al., 
2013).
    Van Woesik et al. (2012) developed a coral resiliency index based 
on biological traits and processes to evaluate extinction risk due to 
bleaching. Evaluations were performed at the genus level, but genera 
were separated between the Caribbean and Indo-Pacific. They indicated 
that A. palmata is highly vulnerable to extinction.
    All information on the susceptibility of A. palmata to ocean 
warming can be summarized as follows. High water temperatures affect A. 
palmata through bleaching, lowered resistance to disease, and effects 
on reproduction. Temperature-induced bleaching and mortality following 
bleaching are temporally and spatially variable. Bleaching associated 
with the high temperatures in 2005 had a large impact on A. palmata 
with 40 to 50 percent of bleached colonies suffering either partial or 
complete mortality in several locations. Algal symbionts did not shift 
in A. palmata after the 1998 bleaching event indicating the ability to 
adapt to rising temperatures may not occur through this mechanism. 
However, Acropora palmata showed evidence of resistance to bleaching 
from warmer temperatures in some portions of its range under some 
circumstances (Little Cayman). Through the effects on reproduction, 
high temperatures can potentially decrease larval supply and settlement 
success, decrease average larval dispersal distances, and cause earlier 
larval settlement, thereby affecting gene flow among populations. 
Therefore, we conclude that A. palmata is highly susceptible to ocean 
warming.
    Information on A. palmata's susceptibility to acidification that we 
considered in the proposed rule includes the following. Ocean 
acidification has a negative impact on early life stages of A. palmata. 
Compared to controls at 400 [mu]atm, carbon dioxide levels of 560 and 
800 [mu]atm, predicted to occur this century, reduced the rate of 
fertilization and settlement (combined 52 and 73 percent, respectively) 
and post-settlement growth (39 and 50 percent, respectively) of A. 
palmata in lab experiments, and impairment of fertilization was 
exacerbated at lower sperm concentrations (Albright et al., 2010).
    The public comments did not provide new or supplemental information 
on the susceptibility of A. palmata to acidification. Supplemental 
information we found on its susceptibility to this threat includes the 
following. No effects on the progression or timing of larval 
development, or embryo and larval size were detected at elevated carbon 
dioxide levels of 700 [micro]atm or 1000 [micro]atm (Medina-Rosas et 
al., 2013).
    All information on the susceptibility of A. palmata to 
acidification can be summarized as follows. Ocean acidification will 
likely impact fertilization, settlement success, and post-settlement 
growth of A. palmata. Therefore, we conclude that A. palmata is highly 
susceptible to acidification.
    There is no species-specific information on the trophic effects of 
fishing on A. palmata. However, due to the level of reef fishing 
conducted in the Caribbean, coupled with Diadema die-off and lack of 
significant recovery,

[[Page 53971]]

recruitment habitat is limited. Therefore, the trophic effects of reef 
fishing adversely affects A. palmata's recruitment habitat. Thus, we 
conclude that A. palmata has some susceptibility to the trophic effects 
of reef fishing due to low natural recruitment rates. However, the 
available information does not support a more precise description of 
susceptibility to this threat.
    Information on A. palmata's susceptibility to sedimentation that we 
considered in the proposed rule includes the following. The morphology 
of A. palmata contributes to its sensitivity to sedimentation as it is 
poorer at removing sediment compared to mounding corals such as 
Orbicella annularis and Diploria strigosa (Abdel-Salam et al., 1988). 
Out of five species tested, A. palmata was the least tolerant of 
sediment exposure; single applications of 200 mg per cm\2\ to colonies 
caused coral tissue death as sediments accumulated on the flattened, 
horizontal surfaces (Rogers, 1983). It is generally unable to remove 
coarser sediments and only weakly able to remove finer sediments 
(Acropora Biological Review Team, 2005). Water movement and gravity are 
probably more important in removing sediments from this species than 
their capabilities of sloughing sediments in stagnant water (Acropora 
Biological Review Team, 2005). Because A. palmata is highly dependent 
on sunlight for nutrition, it is also sensitive to suspended sediments 
that reduce water clarity (Porter, 1976).
    The public comments did not provide new or supplemental information 
on A. palmata's susceptibility to sedimentation. Supplemental 
information we found on the susceptibility of A. palmata to 
sedimentation includes the following. In Vega Baja, Puerto Rico, A. 
palmata mortality increased to 52 percent concurrent with pollution and 
sedimentation associated with raw sewage and beach nourishment, 
respectively, between December 2008 and June 2009 (Hernandez-Delgado et 
al., 2011). Mortality presented as patchy necrosis-like and white pox-
like conditions that impacted local reefs following anthropogenic 
disturbances and was higher inside the shallow platform (52 to 69 
percent) and closer to the source of pollution (81 to 97 percent) 
compared to the outer reef (34 to 37 percent; Hernandez-Delgado et al., 
2011).
    All information on the susceptibility of A. palmata to 
sedimentation can be summarized as follows. Acropora palmata is 
sensitive to sedimentation due to its poor capability of removing 
sediment and its high reliance on clear water for nutrition, and 
sedimentation can cause tissue mortality. We conclude that A. palmata 
is highly susceptible to sedimentation.
    Information on A. palmata's susceptibility to nutrient enrichment 
that we considered in the proposed rule includes the following. There 
are few studies of the effects of nutrients on A. palmata. Field 
experiments indicate that the mean net rate of uptake of nitrate by A. 
palmata exceeds that of ammonium by a factor of two and that A. palmata 
does not uptake nitrite (Bythell, 1990).
    The public comments did not provide new or supplemental information 
on the susceptibility of A. palmata to nutrient enrichment. 
Supplemental information we found on the susceptibility to this threat 
includes the following. In Vega Baja, Puerto Rico, A. palmata mortality 
increased to 52 percent concurrent with pollution and sedimentation 
associated with raw sewage and beach nourishment, respectively, between 
December 2008 and June 2009 (Hernandez-Delgado et al., 2011). Mortality 
presented as patchy necrosis-like and white pox-like conditions that 
impacted local reefs following anthropogenic disturbances and was 
higher inside the shallow platform (52 to 69 percent) and closer to the 
source of pollution (81 to 97 percent) compared to the outer reef (34 
to 37 percent; Hernandez-Delgado et al., 2011).
    All information on the susceptibility of A. palmata to nutrient 
enrichment can be summarized as follows. Acropora palmata is sensitive 
to nutrients as evidenced by increased mortality after exposure to raw 
sewage. We conclude that A. palmata is highly susceptible to nutrient 
enrichment.
    Information on A. palmata's susceptibility to predation that we 
considered in the proposed rule includes the following. There are 
several known predators of A. palmata including the corallivorous snail 
Coralliophila abbreviata (Baums et al., 2003) and the polychaete worm 
Hermodice carrunculata. Incidental corallivores that affect A. palmata 
include several species of fish such as stoplight parrotfish Sparisoma 
viride and three-spot damselfish Stegastes planifrons. Stegastes 
planifrons does not directly feed on the coral but removes live tissue 
to cultivate algal gardens. Likewise, parrotfish are primarily 
herbivores and may be feeding on endolithic algae in coral tissue 
(Bruckner et al., 2000). Monitoring in the Florida Keys indicates that 
parrotfish bites on A. palmata usually heal in a matter of weeks to 
months (Acropora Biological Review Team, 2005). Predators have been 
identified as potential vectors and reservoirs of disease (Sutherland 
et al. 2011).
    The corallivorous snail C. abbreviata is the main predator, 
removing up to 16 cm\2\ of tissue per day (Brawley and Adey 1982), and 
there is evidence that they concentrate on remnant Acropora populations 
following decline (Acropora Biological Review Team, 2005). Severity of 
predation is variable, and Coralliophila seem to be extremely rare or 
absent on Acropora spp. in certain areas such as the Dry Tortugas, 
Florida and Bocas del Toro, Panama (Acropora Biological Review Team, 
2005). In St. John, U.S. Virgin Islands, snail predation affected a 
total of six percent of the colonies across 29 sites, but at individual 
sites, predation affected up to 60 percent of the colonies (Grober-
Dunsmore et al., 2006). In Los Roques, Venezuela snail predation was 
the most common cause of partial mortality (4 to 20 percent), and it 
affected 0.72 to 10.6 percent of the colonies (Zubillaga et al., 2008). 
Surveys of 235 sites throughout the Florida Keys in 2007 revealed that 
about five percent of the A. palmata colonies assessed for condition 
were affected via predation by snails and damselfish (Miller et al., 
2008). In Puerto Rico, infestations of corallivorous snails were 
observed on three percent of all A. palmata colonies surveyed and 
ranged from 0.9 to 10.6 percent per site (Sch[auml]rer et al., 2009).
    The public comments did not provide new or supplemental information 
on the susceptibility of A. palmata to predation. Supplemental 
information we found on the susceptibility of A. palmata to predation 
includes the following. Of the 50 percent tissue loss experienced 
during monitoring in the Florida Keys between 2004 and 2010, snail 
predation accounted for 15 percent after storm damage (42 percent) and 
disease (33 percent; Williams and Miller, 2012). The honeycomb cowfish 
Acanthostracion polygonius has been observed biting A. palmata and 
causing tissue lesions; it is unknown whether the fish is actively 
feeding on the coral tissue or if lesions are a by-product of its 
foraging mode (Williams and Bright, 2013). Lesions healed rapidly (less 
than six weeks) and did not contribute to significant losses of live 
tissue (Williams and Bright, 2013).
    All information on the susceptibility of A. palmata to predation 
can be summarized as follows. Predators can have an impact on A. 
palmata both through tissue removal and the potential to spread 
disease. Predation pressure is spatially variable and almost non-
existent in some locations. However, the

[[Page 53972]]

effects of predation can become more severe if colonies decrease in 
abundance and density, as predators focus on the remaining living 
colonies. Therefore, we conclude that A. palmata has high 
susceptibility to predation.
    Information on A. palmata's susceptibility to sea level rise that 
we considered in the proposed rule includes the following. In-place 
colonies of A. palmata have been used in the geologic record for 
reconstructing Holocene sea level because this species only develops 
monospecific thickets in waters less than 5 m deep and is generally 
limited to depths of 10 m or less (Blanchon, 2005; Blanchon et al., 
2009). A sustained sea level rise of more than 14 mm per year is likely 
to displace A. palmata from its thicket-forming, framework-building 
depth range (<=5 m) into its remaining habitat range where a mixed 
framework is likely to develop (Brainard et al., 2011). In the Yucatan 
region of Mexico during the warming that led to the last interglacial 
period, A. palmata was able to keep up with the first 3 m of rapid sea-
level rise; continued sea-level rise led to the demise of the original 
fore-reef crests inhabited by A. palmata, the retreat of A. palmata to 
a more inland site, and back-stepping of the reef crest as sea level 
rose an additional 2 to 3 m (total of 6 m over an ecological time 
scale; Brainard et al., 2011).
    The public comments did not provide new or supplemental information 
on A. palmata's susceptibility to sea level rise, and we did not find 
any new or supplemental information. All information on the 
susceptibility of A. palmata to sea level rise can be summarized as 
follows. The fast growth rate of A. palmata could accommodate deeper 
water. We conclude that A. palmata has a low susceptibility to sea 
level rise.
    Information on A. palmata's susceptibility to collection and trade 
that we considered in the proposed rule includes the following. Over 
the last decade, collection and trade of this species has been low. The 
public comments did not provide new or supplemental information on the 
susceptibility of A. palmata to collection and trade. Supplemental 
information we found includes the following. Gross exports averaged 
2,120 pieces of coral per year between 2000 and 2012 and have primarily 
been for scientific purposes (data available at http://trade.cites.org). We conclude that A. palmata has low susceptibility to 
collection and trade.
Regulatory Mechanisms
    In the proposed rule, we relied on information from the Final 
Management Report for evaluating the existing regulatory mechanisms for 
controlling threats to all corals. However, we did not provide any 
species-specific information on the regulatory mechanisms or 
conservation efforts for A. palmata. Public comments were critical of 
that approach, and we therefore attempt to analyze regulatory 
mechanisms and conservation efforts on a species basis, where possible, 
in this final rule. We also incorporate here, the evaluation of threats 
to this species conducted in the 2005 status review. Records confirm 
that A. palmata occurs in eight Atlantic ecoregions that encompass 26 
kingdom's and countries' EEZs. The 26 kingdoms and countries are 
Antigua & Barbuda, Bahamas, Barbados, Belize, Colombia, Costa Rica, 
Cuba, Dominica, Dominican Republic, French Antilles, Grenada, 
Guatemala, Haiti, Kingdom of the Netherlands, Honduras, Jamaica, 
Mexico, Nicaragua, Panama, St. Kitts & Nevis, St. Lucia, St. Vincent & 
Grenadines, Trinidad and Tobago, United Kingdom (British Caribbean 
Territories), United States (including U.S. Caribbean Territories), and 
Venezuela. The regulatory mechanisms relevant to A. palmata, described 
first as a percentage of the above kingdoms and countries that utilize 
them to any degree, and second as the percentages of those kingdoms and 
countries whose regulatory mechanisms may be limited in scope, are as 
follows: General coral protection (31 percent with 12 percent limited 
in scope), coral collection (50 percent with 27 percent limited in 
scope), pollution control (31 percent with 15 percent limited in 
scope), fishing regulations on reefs (73 percent with 50 percent 
limited in scope), managing areas for protection and conservation (88 
percent with 31 percent limited in scope). The most common regulatory 
mechanisms in place for A. palmata are fishing regulations and area 
management for protection and conservation. However, half of the 
fishing regulations are limited in scope. General coral protection and 
collection laws, along with pollution control laws, are much less 
common regulatory mechanisms for the management of A. palmata. The 2005 
status review and 2006 listing concluded that existing regulatory 
mechanisms are inadequate to control both global and local threats, and 
are contributing to the threatened status of the species, and we 
incorporate that analysis here.
    Additionally, the public comments suggested that we did not fully 
consider the effects that conservation efforts have on the status of A. 
palmata. Therefore, conservation efforts are described as follows. 
Conservation efforts have been underway for A. palmata for a number of 
years. Of 60 Acropora restoration efforts identified in 14 Caribbean 
countries, 52 percent used A. palmata, including efforts in Belize, 
British Virgin Islands, Colombia, Curacao, Dominican Republic, 
Guadalupe, Jamaica, Mexico, Puerto Rico, Turks and Caicos, U.S. Virgin 
Islands, and Florida (Young et al., 2012). SECORE, a conservation 
organization comprised of public aquariums, zoos, and researchers, 
holds annual workshops to accommodate sexual fertilization of A. 
palmata eggs collected from the wild, with the intent of rearing larvae 
for development of ex situ populations for conservation (Petersen et 
al., 2008). However, to date, A. palmata colonies produced through in 
vitro fertilization have rarely been planted into the wild for 
restoration (but see Roik et al., 2011; Szmant and Miller, 2005).
    Restoration efforts involving A. palmata more typically re-attach 
fragments after physical disturbance such as storms or ship groundings 
(Bruckner and Bruckner, 2001; Garrison and Ward, 2008) or grow colonies 
in coral nurseries (Becker and Mueller, 2001; Bowden-Kerby and Carne, 
2012; Johnson et al., 2011) to outplant. Fast growth rates, branching 
morphology, and asexual reproduction through fragmentation make A. 
palmata an ideal candidate for active propagation, and there are a 
number of offshore nurseries that are producing corals for use in 
restoration and re-establishment of degraded populations. High 
survivorship (>70 percent) of coral fragments has been found within 
coral nurseries during the first year of propagation (Young et al., 
2012). Survival rates after transplanting are variable, ranging between 
43 and 95 percent during the first year, and decreasing in some studies 
using non-nursery raised fragments to 0 to 20 percent after five years 
(Young et al., 2012).
    In conclusion, there are many conservation efforts aimed at 
increasing abundance and genetic diversity of A. palmata throughout the 
Caribbean. These efforts are important, but not enough to ensure 
conservation unless combined with efforts to reduce the underlying 
threats and causes of mortality (Young et al., 2012). While 
conservation efforts will likely enhance recovery and conservation of 
A. palmata at small spatial scales, they are unlikely

[[Page 53973]]

to affect the status of the species, given the global nature of 
threats.
Vulnerability to Extinction
    In 2006, A. palmata was listed as threatened, i.e., likely to 
become in danger of extinction within the next 30 years, due to: (1) 
Recent drastic declines in abundance of the species that have occurred 
throughout its geographic range and abundances at historic lows; (2) 
the potential constriction of broad geographic ranges due to local 
extirpations resulting from a single stochastic event (e.g., 
hurricanes, new disease outbreak); (3) limited sexual recruitment in 
some areas and unknown levels in most; and (4) occurrence of the Allee 
effect (in which fertilization success declines greatly as adult 
density declines).
    The species was not listed as endangered, i.e., currently in danger 
of extinction, because: (1) It was showing limited, localized recovery; 
(2) range-wide, the rate of decline appeared to have stabilized and was 
comparatively slow as evidenced by persistence at reduced abundances 
for the past two decades; (3) it was buffered against major threats by 
the large number of colonies, large geographic range, and asexual 
reproduction; and (4) as shown by the geologic record, the species has 
persisted through climate cooling and heating fluctuation periods over 
millions of years, whereas other corals have gone extinct.
    In 2012, A. palmata was proposed for listing as endangered because 
information available since the original 2006 listing as threatened 
suggested: (1) Population declines have continued to occur, with 
certain populations of both species decreasing up to an additional 50 
percent or more since the time of listing; (2) there are documented 
instances of recruitment failure in some populations; (3) minimal 
levels of thermal stress (e.g., 30 degrees C) have been shown to impair 
larval development, larval survivorship, and settlement success of A. 
palmata; (4) near-future levels of acidification have been demonstrated 
to impair fertilization, settlement success, and post-settlement growth 
rates in A. palmata; (5) on average 50 percent of the colonies are 
clones, meaning the effective number of genetic individuals is half the 
total population size; (6) the species' ranges are not known to have 
contracted, but with continued declines local extirpations are likely, 
resulting in a reduction of absolute range size. Furthermore, we took 
into account that the BRT identified restriction to the Caribbean as a 
spatial factor increasing extinction risk, though, among other things, 
exposure to high levels of human disturbance that result in pollution 
and breakage impacts. Also, while asexual reproduction (fragmentation) 
provides a source for new colonies (albeit clones) that can buffer 
natural demographic and environmental variability remains true, we 
believed that reliance on asexual reproduction is not sufficient to 
prevent extinction of the species. Last, the previous status review and 
listing determination underestimated the global climate change-
associated impacts to A. palmata and A. cervicornis, based on our 
current knowledge of trends in emissions, likely warming scenarios, and 
ocean acidification. In particular, in the previous determination, we 
identified ocean acidification only as a factor that ``may be 
contributing'' to the status of two species, in comparison to our 
current understanding that ocean acidification is one of the three 
highest order threats affecting extinction risk for corals.
    As explained above in the Risk Analyses section, a species' 
vulnerability to extinction results from the combination of its spatial 
and demographic traits, threat susceptibilities, and consideration of 
the baseline environment and future projections of threats. Subsequent 
to the proposed rule, we received and gathered supplemental species- or 
genus-specific information, described above, that expands our knowledge 
regarding the species' abundance, distribution, and threat 
susceptibilities. We developed our assessment of the species' 
vulnerability to extinction using all the available information. As 
explained in the Risk Analyses section, our assessment in this final 
rule emphasizes the ability of the species' spatial and demographic 
traits to moderate or exacerbate its vulnerability to extinction, as 
opposed to the approach we used in the proposed rule, which emphasized 
the species' susceptibility to threats.
    The following characteristics of A. palmata, in conjunction with 
the information described in the Corals and Coral Reefs section, Coral 
Habitat sub-section, and Threats Evaluation section above, affect its 
vulnerability to extinction currently and over the foreseeable future. 
The species has undergone substantial population decline and decreases 
in the extent of occurrence throughout its range due mostly to disease. 
Although localized mortality events have continued to occur, percent 
benthic cover and proportion of reefs where A. palmata is dominant have 
remained stable over its range since the mid-1980s. There is evidence 
of synergistic effects of threats for this species including disease 
outbreaks following bleaching events. Acropora palmata is highly 
susceptible to a number of threats, and cumulative effects of multiple 
threats are likely to exacerbate vulnerability to extinction. Despite 
the large number of islands and environments that are included in the 
species' range, geographic distribution in the highly disturbed 
Caribbean exacerbates vulnerability to extinction over the foreseeable 
future because A. palmata is limited to an area with high localized 
human impacts and predicted increasing threats. Acropora palmata occurs 
in turbulent water on the back reef, fore reef, reef crest, and spur 
and groove zone in water ranging from 1 to 30 m in depth. This 
moderates vulnerability to extinction over the foreseeable future 
because the species occurs in numerous types of reef environments that 
will, on local and regional scales, experience highly variable thermal 
regimes and ocean chemistry at any given point in time. Its absolute 
population abundance has been estimated as at least hundreds of 
thousands of colonies in both Florida and a portion of the U.S. Virgin 
Islands and is higher than the estimate from these two locations due to 
the occurrence of the species in many other areas throughout its range. 
Acropora palmata has low sexual recruitment rates, which exacerbates 
vulnerability to extinction due to decreased ability to recover from 
mortality events when all colonies at a site are extirpated. In 
contrast, its fast growth rates and propensity for formation of clones 
through asexual fragmentation enables it to expand between rare events 
of sexual recruitment and increases its potential for local recovery 
from mortality events, thus moderating vulnerability to extinction. Its 
abundance and life history characteristics, combined with spatial 
variability in ocean warming and acidification across the species' 
range, moderate vulnerability to extinction because the threats are 
non-uniform, and there will likely be a large number of colonies that 
are either not exposed or do not negatively respond to a threat at any 
given point in time.
Listing Determination
    In the proposed rule, using the determination tool formula 
approach, A. palmata was proposed for listing as endangered because of: 
High vulnerability to ocean warming (E), ocean acidification (E) and 
disease (C); high vulnerability to sedimentation (A and E) and nutrient 
over-enrichment (A and E); uncommon abundance (E); decreasing trend in 
abundance (E); low

[[Page 53974]]

relative recruitment rate (E); narrow overall distribution (E); 
restriction to the Caribbean (E); and inadequacy of regulatory 
mechanisms (D).
    In this final rule, we changed the listing determination for A. 
palmata from endangered to threatened. We made this determination based 
on a more species-specific and holistic approach, including 
consideration of the buffering capacity of this species' spatial and 
demographic traits, and the best available information above on A. 
palmata's spatial structure, demography, threat susceptibilities, and 
management. The combination of these factors indicates that A. palmata 
is likely to become endangered throughout its range within the 
foreseeable future, and thus warrants listing as threatened at this 
time, because:
    (1) Acropora palmata is highly susceptible to ocean warming (ESA 
Factor E), disease (C), ocean acidification (E), sedimentation (A, E), 
nutrients (A, E), and predation (C) and susceptible to trophic effects 
of fishing (A), depensatory population effects from rapid, drastic 
declines and low sexual recruitment (C), and anthropogenic and natural 
abrasion and breakage (A, E). These threats are expected to continue 
and increase into the future. In addition, the species is at heightened 
extinction risk due to inadequate existing regulatory mechanisms to 
address local and global threats (D);
    (2) Acropora palmata is geographically located in the highly 
disturbed Caribbean, where localized human impacts are high and threats 
are predicted to increase as described in the Threats Evaluation 
section. A range constrained to this particular geographic area that is 
likely to experience severe and increasing threats indicates that a 
high proportion of the population of this species is likely to be 
exposed to those threats over the foreseeable future; and
    (3) Acropora palmata's abundance is still a fraction of what it was 
before the mass mortality in the 1970s and 1980s, and recent population 
models forecast the extirpation of the species from some locations over 
the foreseeable future.
    The combination of these characteristics and future projections of 
threats indicates that the species is likely to be in danger of 
extinction within the foreseeable future throughout its range and 
warrants listing as threatened at this time due to factors A, C, D, and 
E.
    The available information above on A. palmata's spatial structure, 
demography, threat susceptibilities, and management also indicate that 
the species is not currently in danger of extinction and thus does not 
warrant listing as Endangered because:
    (1) While A. palmata's distribution in the Caribbean increases its 
risk of exposure to threats as described above, its habitat includes 
back reef environments and turbulent water on the fore reef, reef 
crest, shallow spur and groove zone. It is most commonly found in 
depths of one to 12 m but is also found in depths up to 30 m. This 
moderates vulnerability to extinction currently because the species is 
not limited to one habitat type but occurs in numerous types of reef 
environments that will experience highly variable thermal regimes and 
ocean chemistry on local and regional scales at any given point in 
time, as described in more detail in the Coral Habitat and Threats 
Evaluation sections;
    (2) Acropora palmata's absolute abundance is at least hundreds of 
thousands of colonies based on estimates from two locations in its 
range. Absolute abundance is higher than estimates from these locations 
since A. palmata occurs in many other locations throughout its range. 
This absolute abundance allows for variation in the responses of 
individuals to threats to play a role in moderating vulnerability to 
extinction for the species to some degree, as described in more detail 
in the Corals and Coral Reefs section;
    (3) Recent information indicates that proportions of Caribbean 
sites where A. palmata is present and dominant have stabilized;
    (4) Acropora palmata has fast growth rates and high capacity to 
produce clones through asexual fragmentation, which can aid in local 
recovery from mortality events; and
    (5) Acropora palmata shows evidence of resistance to bleaching from 
warmer temperatures in some portions of its range under some 
circumstances (e.g. Little Cayman).
    The combination of these characteristics indicates that the species 
does not exhibit the characteristics of one that is currently in danger 
of extinction, as described previously in the Risk Analyses section and 
thus does not warrant listing as endangered at this time. Therefore, we 
withdraw our proposal to list A. palmata as endangered.
    Progress has been made with A. palmata-specific conservation and 
restoration projects, albeit small-scale, and these projects are likely 
to increase in the future. Within some countries, A. palmata-specific 
conservation and restoration projects show promise for enhancing 
species recovery at very small spatial scales and facilitating the 
persistence of the species in some areas in the face of continuing 
threats. Range-wide, a multitude of conservation efforts are already 
broadly employed specifically for A. palmata. However, considering the 
global scale of the most important threats to the species, and the 
ineffectiveness of conservation efforts at addressing the root cause of 
global threats (i.e., GHG emissions), we do not believe that any 
current conservation efforts or conservation efforts planned in the 
future will result in affecting the species' status to the point at 
which listing is not warranted.

Indo-Pacific Species Determinations

    Absolute abundance is approximated at a coarse scale in the 
Demographic Information sections for most of the Indo-Pacific species, 
based on a comparison of corrected data from Richards et al. (2008) and 
the distribution and abundance results from Veron (2014). Mean global 
census sizes for four species in this final rule (Acropora 
jacquelineae, A. lokani, A. speciosa, and A. tenella) are provided in 
Richards et al. (2008). An error in the global census size formula 
(Richards et al. 2008, Supplementary Information file 
MethodsS1) resulted in 1,000-fold under-estimates of global 
census size in Richards et al. (2008) for these four species, as 
confirmed by NMFS with the author in 2013. Richards et al.'s (2008) 
corrected census results were compared with Veron's ecoregion 
distribution and semi-quantitative abundance results to derive coarse 
approximations of absolute abundance. For each species, the resulting 
absolute abundance is described as either ``at least millions of 
colonies,'' or ``at least tens of millions of colonies'' (NMFS, 2014). 
Although this comparison produces only very general approximations of 
abundance, large scale estimates are sufficient for considering whether 
population size provides buffering capacity within the context of our 
listing determinations.

Genus Millepora

Genus Introduction
    The SRR and SIR provided no genus-level introduction information 
for Millepora. However, they did provide the following information on 
reproduction in the genus. Millepora species are hydrozoans, thus their 
life history cycle includes a medusae stage, a free-swimming, bell-
shaped form (``jellyfish'') that produces gametes. Reproduction is 
seasonal. The adult coral colonies produce tiny medusae, which release 
gametes within a few days after being released from the colony. Medusae 
are in separate sexes,

[[Page 53975]]

and the milleporid medusae of some species live for only a few hours. 
The gametes of some milleporids can become mature in 20 to 30 days, 
more rapidly than for many scleractinians. Hydrozoan corals of the 
genus Millepora are the only reef-building corals with medusae as part 
of their life history. Branching and columnar forms of Millepora are 
subject to fragmentation and may use this mechanism to reproduce 
asexually; unlike scleractinian corals, the survival of Millepora 
fragments may not be size-dependent.
    There is only one genus in the Family Milleporidae, the genus 
Millepora. About 16 species of Millepora are currently considered 
valid. While all coral species in this final rule are ``cnidarians'' 
(Phylum Cnidaria), Millepora are ``hydrozoans'' (Class Hydrozoa, which 
includes jellyfish), whereas all other species in this rule are 
``scleractinians'' (Class Anthozoa, Order Scleractinia). Like other 
reef-building corals, Millepora species contain zooxanthellae, produce 
calcium carbonate skeletons, may grow fast, and are thus major 
contributors to the physical structure of coral reefs. Unlike other 
reef-building corals, the surfaces of Millepora colonies are covered 
with tiny polyps that look like hairs, containing stinging cells to 
capture prey. Most species can sting humans with the same stinging 
cells, hence the common name ``fire corals.'' Colonies of Millepora 
species are encrusting, branching, foliose (leafy), or combinations of 
these forms. The biology and ecology of Millepora are reviewed in Lewis 
(2006).
Genus Susceptibility to Threats
    The SRR and SIR provided the following information on the threat 
susceptibilities of the genus Millepora. The genus Millepora has been 
called a bleaching ``loser.'' Millepora species are ranked as the most 
susceptible to bleaching in response to high seawater temperatures of 
any of the 40 genera or other categories of hermatypic corals in the 
Great Barrier Reef. The genus has been reported to be highly 
susceptible to bleaching in the western Indian Ocean and appears to 
have experienced local extirpations in the tropical eastern Pacific. 
Low bleaching occurred in Millepora in Moorea during the 1991 event, 
but elevated temperatures can also kill Millepora even in the absence 
of bleaching. At elevated temperatures, Millepora dichotoma showed 
decreased zooxanthellae density, changes in chlorophyll concentrations, 
and decreased calcification. Millepora species are among the first to 
bleach and die in response to high temperature events, but they also 
seem to have a high capacity for quickly recovering by recruiting new 
colonies.
    Millepora have been observed with a greater than 20 percent 
prevalence of skeleton-eroding-band disease in the Red Sea. There are 
reports of black-band disease on Millepora on the Great Barrier Reef 
and white plague in Florida. Few other reports exist for the Pacific, 
and Caribbean congeners have been observed with a small number of 
diseases.
    Millepora species are known to be preyed on by the crown-of-thorns 
seastar Acanthaster planci, although they are less preferred prey than 
acroporids and perhaps most scleractinians. Millepora species are also 
preyed on by the polychaete Hermodice carunculata, the nudibranch 
mollusk Phyllidia, and filefish of the genera Alutera and Cantherhines.
    Although Millepora species tend to favor relatively clear water 
with low rates of sedimentation, they were reported to be among the 
last 17 out of 42 genera to drop out along a gradient of increasing 
rate of sedimentation. Millepora also showed increased relative 
abundance and colony size on sediment impacted reefs in Kenya. Though 
little is known about effects of nutrients on Pacific Millepora, 
Caribbean congeners were found to decrease in percent cover on 
eutrophic reefs in Barbados.
    The genus Millepora has been involved in international trade from 
Indonesia, Solomon Islands, and Fiji with reported exports between 200 
and 3000 pieces per year in the years 2000-2008. Reported exports from 
Vietnam, Malaysia, and Tonga were less than 1000 pieces per year in the 
same time period.
    Public comments did not provide any information on the genus 
Millepora. We gathered supplemental information on the susceptibilities 
of Millepora species to some threats, including the following. High 
bleaching and mortality in Millepora species has been reported in 
response to warming events. All Millepora colonies on reef flats of two 
islands in the Thousand Islands of Indonesia died in the 1983 El Nino 
mass bleaching (Brown and Suharsono, 1990). In contrast, Millepora 
colonies showed no evidence of bleaching in Moorea, French Polynesia in 
the 1991 bleaching event other than occasional mild paling (Gleason, 
1993). In Palau in 2000, some mortality was seen among Millepora 
colonies (Bruno et al., 2001). Almost all Millepora colonies in study 
sites outside of marine protected areas in Kenya were killed by mass 
bleaching in 1998, but in protected sites there was actually an 
increase in Millepora colonies (McClanahan et al., 2001). Millepora 
colonies had a bleaching index of 23.06 for eight countries in the 
western Indian Ocean in 1998-2005, which was 12th highest of the 45 
genera recorded, and 56 percent of the highest value (McClanahan et 
al., 2007a). Millepora had the highest bleaching level of any genus in 
Australia, and a moderately high level in Kenya in 1998 (Pandolfi et 
al., 2011). Millepora colonies in Okinawa, Japan, experienced sharp 
drops in populations following the 1998 and 2010 mass bleaching 
episodes (Hongo and Yamano, 2013). At Mauritius in a bleaching event in 
2004, Millepora had a bleaching index of 35, the second highest of the 
32 genera recorded (McClanahan et al., 2005a). Millepora colonies had 
the highest level of bleaching among the corals of the Socotra islands 
of Yemen, just outside the Red Sea, in 1998 (DeVantier et al., 2005).
    While Millepora species are among the most susceptible of all reef-
building corals to warming-induced bleaching, they also often recover 
more quickly than scleractinians, opportunistically over-growing 
bleached colonies. Such relatively rapid recovery by Millepora species 
from bleaching events has been observed in both the Indo-Pacific and 
Caribbean, and is facilitated by short colony life and ready 
regeneration of fragments (Lewis, 2006). At a forereef site in the 
Marquesas Islands, French Polynesia, Millepora platyphyllia overgrew 
dead scleractinian colonies to form a large monospecific stand 
(Andr[eacute]fou[euml]t et al., 2014). At a back-reef site on Ofu 
Island, American Samoa, following a bleaching event in 2002 that killed 
almost all Millepora dichotoma, colonies appeared and became fairly 
common within a few years (Doug Fenner, personal comm.). Following both 
the 1982-83 and 1997-98 warming events, Millepora intricata was 
extirpated from shallow water in the eastern Pacific, but showed 
recovery within several years, likely because of recolonization from 
deep water (Smith et al., in press). In contrast, a long-term study 
showed that three Millepora species were ``long-term losers'' (i.e., 
populations decreased to zero, and remained there) following mass 
bleaching events in Japan in 1998 and 2010, while two other species of 
Millepora were ``neither winners nor losers'' (i.e., changes in their 
populations were not significant) (van Woesik et al., 2011).
    Millepora colonies in the Great Barrier Reef had low susceptibility 
to Skeletal Eroding Band (the most prevalent

[[Page 53976]]

disease on the GBR), with a prevalence of 0.4 percent out of 4,068 
colonies surveyed (Page and Willis, 2007).
    Several recent studies compare vulnerabilities across genera or 
species for a large number of reef-building coral species, and the 
results of these studies are summarized below with regard to Millepora. 
Foden et al. (2013) developed a framework for evaluating the 
vulnerability of corals (and birds and amphibians) to extinction due to 
climate change. They categorized all of the six species of Millepora, 
which they reported on as having a low vulnerability to climate change. 
A field study that tracked the responses of 46 reef-building coral 
species in southern Japan from 1997 to 2010 through two bleaching 
events in 1998 and 2001 rated three Millepora species as neither 
winners nor losers, and two Millepora species as long term losers. 
Three of the Millepora species were branching and all three branching 
species were ``long term losers,'' one species is encrusting and one 
produces plates, and those two species were neither long term winners 
or losers (van Woesik et al., 2011). There is no information available 
on the effects of any other threat for Millepora species.
Genus Conclusion
    Based on the information from the SRR, SIR, public comments, and 
supplemental information we can make the following inferences about the 
susceptibilities of an unstudied Millepora species to ocean warming, 
disease, ocean acidification, trophic effects of fishing, 
sedimentation, nutrients, sea-level rise, predation, or collection and 
trade. The large majority of studies report that Millepora species are 
highly susceptible to thermal stress and bleaching, but vulnerability 
may be moderated by the capacity for rapid recovery in some species. An 
unstudied species of Millepora such as M. tuberosa can be predicted in 
a bleaching event to not be a ``winner'' in the long term, but it 
cannot be predicted whether they will be a long term loser, or neither 
a winner or loser. Thus, an unstudied species of Millepora is likely to 
be highly susceptible to ocean warming. Based on the above information, 
an unstudied species of Millepora is likely to have some susceptibility 
to disease, sedimentation, nutrients, and predation.
    The SRR rated ocean acidification as ``medium-high'' importance, 
the third most important threat to corals overall, because of the 
likely effects of decreasing ocean pH on coral calcification and 
reproduction. Thus, an unstudied Millepora species is likely to have 
some susceptibility to ocean acidification. The SRR rated the trophic 
effects of fishing as ``medium'' importance, the fourth most important 
threat to corals overall. This threat was not addressed at the genus or 
species level in the SRR or SIR, because it is an ecosystem-level 
process. That is, removal of herbivorous fish from coral reef systems 
by fishing alters trophic interactions by reducing herbivory on algae, 
thereby providing a competitive advantage for space to algae over 
coral. Thus, the SRR did not discuss this threat in terms of coral 
taxa, as its effects are difficult to distinguish between coral genera 
and species. Therefore, an unstudied Millepora species is likely to 
have some susceptibility to the trophic effects of fishing. The SRR 
rated sea-level rise as ``low-medium'' importance to corals overall. 
This threat was not addressed at the genus or species level in the SRR 
or SIR. Increasing sea levels may increase land-based sources of 
pollution due to inundation, resulting in changes to coral community 
structure, most likely to sediment-tolerant assemblages and slower 
growing species. Because Millepora are not generally sediment-tolerant 
and are faster growing species, an unstudied Millepora species is 
likely to have some susceptibility to sea-level rise. The SRR rated 
ornamental trade (referred to in the proposed rule as Collection and 
Trade) as ``low'' importance to corals overall, and this threat is 
addressed at both the genus and species levels in the SRR. Because 
Millepora species are widely collected and traded, an unstudied 
Millepora species is likely to have some susceptibility to collection 
and trade.
    In conclusion, an unstudied Millepora species is likely to be 
highly susceptible to ocean warming (i.e., thermal stress, leading to 
warming-induced bleaching), and to have some susceptibility to disease, 
ocean acidification, trophic effects of fishing, sedimentation, 
nutrients, sea-level rise, predation, and collection and trade.

Millepora foveolata

Introduction
    The SRR and SIR provided the following information on M. 
foveoloata's morphology and taxonomy. Colonies of Millepora foveolata 
form thin encrusting laminae that adhere closely to the underlying 
substrata. Millepora foveolata is sometimes confused with the similarly 
encrusting Millepora exaesa.
    The public comments did not provide any new or supplemental 
information on M. foveoloata's morphology and taxonomy. However, we 
gathered supplemental information on M. foveoloata that indicates a 
very high level of species identification uncertainty, because its 
distinctive features are very small and difficult to learn. In 
addition, no pictures of live colonies have been published of this 
species. Corals of the World (Veron, 2000) does not include non-
scleractinians such Millepora species, making it very difficult to 
obtain reliable reference material. Many coral experts also ignore 
Millepora species, but even those that are interested in them have 
little opportunity to hone identification skills because the species is 
quite rare and not often encountered on surveys. Thus, even though M. 
foveolata is considered a valid species, and there are no known 
taxonomic uncertainty issues, the species is so difficult to identify 
in the field that there is very little reliable information available 
for this species (Fenner, 2014b). Thus, a high proportion of the 
information on M. foveolata's distribution and abundance information in 
the SRR or SIR is likely based on inaccurate field identifications, 
thus we do not consider this information to be sufficiently reliable, 
and are unable to provide a reliable species description for M. 
foveolata in this final rule.
Listing Determination
    In the proposed rule, M. foveolata was proposed for listing as 
endangered because of: High vulnerability to ocean warming (ESA Factor 
E); moderate vulnerability to disease (C) and acidification (E); 
uncommon generalized range wide abundance (E); narrow overall 
distribution (based on narrow geographic distribution and shallow depth 
distribution (E); and inadequacy of existing regulatory mechanisms (D).
    Based on the lack of information on M. foveolata's distribution, 
abundance, and threat vulnerabilities due to this species' 
identification uncertainty, we believe there is not sufficient evidence 
to support a listing determination of threatened or endangered. 
Therefore, we find that listing is not warranted at this time.

Millepora tuberosa

Introduction
    The SRR and SIR provided the following information on M. tuberosa's 
morphology and taxonomy. Millepora tuberosa's colony morphology 
consists of thin (about 1 mm at encrusting peripheral margins) to 
moderately thick (3 cm or more in the central regions of larger 
colonies) encrusting laminae that closely adhere to the underlying 
substrata. They are always encrusting

[[Page 53977]]

and so do not make vertical plates or branches, although they can be 
nodular or lumpy, especially when they encrust rubble. Millepora 
tuberosa is often found as small colonies (5 to 30 cm diameter) but can 
be greater than one meter in diameter. The SIR reports that several 
authors have commented that people could inadvertently misidentify M. 
tuberosa colonies as crustose coralline algae, and the SIR reports it 
can look similar to Psammocora nierstrazi if they have similar color. 
There is some taxonomic uncertainty, as M. tuberosa has been 
synonymized with Millepora exaesa in one review. The problem may be 
that the skeletons are quite similar, but the living colonies appear 
quite different, mainly in color; M. tuberosa is a wine color, unlike 
other Millepora species.
    The public comments and information we gathered did not provide any 
new or supplemental information on morphology or taxonomy. We gathered 
supplemental information, which confirmed that M. tuberosa has moderate 
taxonomic uncertainty, but is easily identified. Millepora tuberosa is 
distinctive and not difficult to identify by experts, thus the 
distribution and abundance information described below for this species 
is sufficiently reliable (Fenner, 2014b).
 Spatial Information
    The SRR and SIR provided the following information on M. tuberosa's 
distribution, habitat, and depth range. Millepora tuberosa is known 
from Mauritius, Taiwan, Mariana Islands, Caroline Islands, American 
Samoa, and New Caledonia. The species occurs in a broad range of 
habitats on the reef slope, reef crest, and back-reef, including but 
not limited to lower reef crests, upper reef slopes, and lagoons, from 
1 to at least 12 m depth.
    Public comments and information we gathered provided new or 
supplemental information on M. tuberosa's distribution. One public 
comment stated M. tuberosa has been reported from a variety of sources 
suggesting that its range extends from that shown in the proposed rule 
westward to Madagascar, indicating a broader distribution than shown in 
the proposed rule. We gathered supplemental information, including 
results from surveys carried out from 2005 to 2014 in New Caledonia, 
American Samoa, the Northern Mariana Islands, Nauru, Tonga, and the 
Chagos Islands, that confirmed the occurrence of M. tuberosa in the 
first three areas but did not find it in the latter three areas (D. 
Fenner, personal comm.). Many experts, including Veron, do not record 
the presence of Millepora species, thus the small number of reliable 
observations for this species likely indicates under-reporting rather 
than a reflection of its actual distribution or overall abundance. 
However, surveys by Millepora experts have not found the species at all 
coral reef sites surveyed within the areas encompassed by its known 
locations. Thus we conclude that the available information suggests a 
patchy range bounded by east Africa, Taiwan, Mariana Islands, Caroline 
Islands, American Samoa, and New Caledonia, and that the species' range 
makes up approximately one third to one half of the coral reef areas 
within the Indo-Pacific.
Demographic Information
    The SRR and SIR provided the following information on M. tuberosa's 
abundance. The SRR stated that the species is most often reported as 
occasional, but in Guam it is predominant in an area of lagoonal reef 
south of Agat Boat Harbor. The SIR cited several sources of information 
not available in the SRR, and concluded that the species' abundance 
should be considered common.
    The public comments did not provide any new or supplemental 
information on M. tuberosa's abundance. We gathered supplemental 
information, including abundance results from surveys conducted in New 
Caledonia, American Samoa, and the Northern Mariana Islands between 
2005 and 2013. In New Caledonia, 87 sites were surveyed from 2006 to 
2009, and only a single colony of M tuberosa was found. At 67 sites 
surveyed in American Samoa from 2005 to 2010, M. tuberosa was found at 
18 sites (of the sites, 31 were on Tutuila, and the species was found 
at 13 of them). At 22 sites surveyed in the Northern Mariana Islands in 
2013, M. tuberosa was found at three sites (D. Fenner, personal comm.). 
At sites where M. tuberosa has been actively surveyed (i.e., by coral 
abundance monitoring programs that includes Millepora experts), the 
available information shows wide variability in the species' abundance, 
from dominant or common (Guam) to uncommon (Tutuila, Northern Mariana 
Islands) to rare (New Caledonia). Based on the available information, 
we conclude that M. tuberosa's overall abundance is common or uncommon 
overall, but locally rare.
    Carpenter et al. (2008) extrapolated species abundance trend 
estimates from total live coral cover trends and habitat types. For M. 
tuberosa, the overall decline in abundance (``Percent Population 
Reduction'') was estimated at 59 percent, and the decline in abundance 
before the 1998 bleaching event (``Back-cast Percent Population 
Reduction'') was estimated at 22 percent (Carpenter et al., 2008). This 
estimated decline is approximately 50 percent higher than most other 
Indo-Pacific species included in the Carpenter paper, apparently 
because of the combined restricted geographic and depth ranges. 
However, as summarized above in the Inter-basin Comparison sub-section, 
live coral cover trends are highly variable both spatially and 
temporally, producing patterns on small scales that can be easily taken 
out of context, thus quantitative inferences to species-specific trends 
should be interpreted with caution. At the same time, an extensive body 
of literature documents broad declines in live coral cover and shifts 
to reef communities dominated by hardier coral species or algae over 
the past 50 to 100 years (Birkeland, 2004; Fenner, 2012; Pandolfi et 
al., 2003; Sale and Szmant, 2012). These changes have likely occurred, 
and are occurring, from a combination of global and local threats. 
Given that M. tuberosa probably occurs in many areas affected by these 
broad changes, and that it has some susceptibility to both global and 
local threats, we conclude that it is likely to have declined in 
abundance over the past 50 to 100 years, but a precise quantification 
is not possible based on the limited species-specific information.
Other Biological Information
    The public comments and information we gathered did not provide 
additional biological information on M. tuberosa.
Susceptibility to Threats
    The SRR and SIR provided species-specific information on the 
susceptibility of M. tuberosa to sedimentation, predation, and 
secondary effects of heavy fishing pressure. The relatively high 
abundance of this species on Guam suggests it is resistant to those 
threats. Genus-level information is provided for the effects on 
Millepora of ocean warming, disease, predation, land-based sources of 
pollution (i.e., sedimentation, nutrients, toxins, and salinity), and 
collection and trade. The SRR and SIR did not provide any other 
species-specific information on the effects of these threats on M. 
tuberosa. The threat susceptibility and exposure information from the 
SRR and SIR was interpreted in the proposed rule for M. tuberosa's 
vulnerabilities to threats as follows: High vulnerability to ocean 
warming, moderate vulnerabilities to disease, acidification, trophic 
effects of fishing, nutrients, and

[[Page 53978]]

low vulnerabilities to predation, sedimentation, sea-level rise, and 
collection and trade.
    Public comments did not provide any new or supplemental information 
on M. tuberosa's threat susceptibilities. We gathered the following 
species-specific and genus-level supplemental information on this 
species' threat susceptibilities. Millepora tuberosa has been rated as 
moderately or highly susceptible to bleaching but not coral disease, 
but these ratings are not based on species-specific data (Carpenter et 
al., 2008). Some colonies in American Samoa and Guam have been observed 
to have a discolored yellow area around part of the perimeter, which 
appeared to be a non-lethal disease (not all colonies had it, and no 
mortality was seen. No other disease was seen (Fenner, 2014a). There is 
no other species-specific information for the exposure or 
susceptibility of M. tuberosa to any threat. Based on information 
provided in the genus description above, M. tuberosa is likely to be 
highly susceptible to ocean warming, and has some susceptibilities to 
disease, ocean acidification, trophic effects of fishing, 
sedimentation, nutrients, sea-level rise, predation, and collection and 
trade.
Regulatory Mechanisms
    In the proposed rule, we did not provide any species-specific 
information on regulatory mechanisms or conservation efforts for M. 
tuberosa. Criticisms of our approach received during public comment led 
us to the following analysis to attempt to analyze regulatory 
mechanisms on a species basis.
    Veron's updated report on the listed coral species and their 
occurrence in various ecoregions (Veron, 2014) did not include M. 
tuberosa. To determine what countries the species occurs in we used the 
SRR, IUCN Red List of Threatened Species, and other sources where the 
species has been confirmed (Fenner, 2011) and conclude that the species 
occurs in a minimum of six countries' EEZs. Those six countries are the 
Federated States of Micronesia, France (New Caledonia), Mauritius, 
Palau, Taiwan, and the United States (CNMI, Guam, American Samoa). As 
noted in the Spatial Information paragraph above, it is likely the 
species occurs in a number of other countries, but we cannot determine 
which ones at this time, thus this management analysis is limited to 
the six countries where the species has been confirmed.
    The regulatory mechanisms available to M. tuberosa, described first 
as a percentage of the above countries that utilize them to any degree, 
and second as the percentage of those countries whose regulatory 
mechanisms are limited in scope, are as follows: General coral 
protection (33 percent with none limited in scope), coral collection 
(67 percent with 17 limited in scope), pollution control (33 percent 
with 17 percent limited in scope), fishing regulations on reefs (100 
percent with 17 percent limited in scope), managing areas for 
protection and conservation (100 percent with none limited in scope). 
The most common regulatory mechanisms in place for M. tuberosa are reef 
fishing regulations and area management for protection and 
conservation. Coral collection laws are also somewhat utilized. General 
coral protection and pollution control laws are much less common 
regulatory mechanisms for the management of M. tuberosa.
Vulnerability to Extinction
    As explained above in the Risk Analyses section, a species' 
vulnerability to extinction results from the combination of its spatial 
and demographic characteristics, threat susceptibilities, and 
consideration of the baseline environment and future projections of 
threats. The SRR stated that the high bleaching rate, based on genus-
level information, is the primary threat of extinction for M. tuberosa, 
which was compounded by the disjunct geographic range. The SRR also 
stated that factors that potentially reduce the extinction risk are 
that M. tuberosa might be more common than previously observed, and 
that like other Millepora species, it likely has a high capacity for 
recovering from bleaching.
    Subsequent to the proposed rule, we received and gathered 
supplemental species- or genus-specific information, described above, 
that expands our knowledge regarding the species abundance, 
distribution, and threat susceptibilities. We developed our assessment 
of the species' vulnerability to extinction using all the available 
information. As explained in the Risk Analyses section, our assessment 
in this final rule emphasizes the ability of the species' spatial and 
demographic traits to moderate or exacerbate its vulnerability to 
extinction, as opposed to the approach we used in the proposed rule, 
which emphasized the species' susceptibility to threats.
    The following characteristics of M. tuberosa, in conjunction with 
the information described in the Corals and Coral Reefs section, Coral 
Habitat sub-section, and Threats Evaluation section above, affect its 
vulnerability to extinction currently and over the foreseeable future. 
Its geographic distribution, based on the available information, 
includes patchy areas from the western Indian Ocean across the western 
and central Pacific, as far east as American Samoa. Its geographic 
distribution moderates vulnerability to extinction because some areas 
within its range are projected to have less than average warming and 
acidification over the foreseeable future, including the western Indian 
Ocean, the central Pacific, and other areas, so portions of the 
population in these areas will be less exposed to severe conditions. 
Its depth range is from zero to at least 12 meters. On one hand, its 
depth range may moderate vulnerability to extinction over the 
foreseeable future because deeper areas of its range will usually have 
lower irradiance than surface waters, and acidification is generally 
predicted to accelerate most in waters that are deeper and cooler than 
those in which the species occurs. On the other hand, its depth range 
may exacerbate vulnerability to extinction over the foreseeable future 
if the species occurs predominantly in the shallower portion of its 
depth range, since those areas will have higher irradiance and thus be 
more severely affected by warming-induced bleaching. Its habitat 
includes lower reef crests, upper reef slopes, and lagoons, which 
moderates vulnerability to extinction over the foreseeable future 
because the species is not limited to one habitat type but occurs in 
numerous types of reef environments that will, on local and regional 
scales, experience reef environments that will, on local and regional 
scales, experience highly variable thermal regimes and ocean chemistry 
at any given point in time. While the species is locally rare, its 
overall abundance is common or uncommon. Thus, its overall abundance, 
combined with spatial variability in ocean warming and acidification 
across the species range, moderates vulnerability to extinction because 
the increasingly severe conditions expected in the foreseeable future 
will be non-uniform and therefore will likely be a large number of 
colonies that are either not exposed or do not negatively respond to a 
threat at any given point in time.
Listing Determination
    In the proposed rule using the determination tool formula approach, 
M. tuberosa was proposed for listing as threatened because of: High 
vulnerability to ocean warming (ESA Factor E); moderate vulnerability 
to disease (C) and acidification (E); common generalized range wide 
abundance (E); narrow overall

[[Page 53979]]

distribution (based on narrow geographic distribution and shallow depth 
distribution (E); and inadequacy of existing regulatory mechanisms (D).
    In this final rule, we changed the listing determination for M. 
tuberosa from threatened to not warranted. We made this determination 
based on a more species-specific and holistic assessment of whether 
this species meets the definition of either a threatened or endangered 
coral largely in response to public comments, including more 
appropriate consideration of the buffering capacity of this species' 
spatial and demographic traits to lessen its vulnerability to threats. 
Thus, based on the best available information above on M. tuberosa's 
spatial structure, demography, threat susceptibilities, and management, 
none of the five ESA listing factors, alone or in combination, are 
causing this species to be likely to become endangered throughout its 
range within the foreseeable future, and thus it is not warranted for 
listing at this time, because:
    (1) Millepora tuberosa's distribution stretches across the Indian 
Ocean and most of the Pacific Ocean and is spread over a very large 
area. While some areas within its range are projected to be affected by 
warming and acidification, other areas are projected to have less than 
average warming and acidification, including the western Indian Ocean, 
the central Pacific, and other areas. This distribution and the 
heterogeneous habitats it occupies reduce exposure to any given threat 
event or adverse condition that does not occur uniformly throughout the 
species range. As explained above in the Threats Evaluation section, we 
have not identified any threat that is expected to occur uniformly 
throughout the species range within the foreseeable future; and
    (2) Millepora tuberosa's abundance is described as common or 
uncommon overall which, in terms of relative abundance of corals and in 
combination with the size of its range, indicates this species likely 
numbers in the tens or hundreds of millions of colonies, at least. This 
provides buffering capacity in the form of absolute numbers of colonies 
and variation in susceptibility between individual colonies. As 
discussed in the Corals and Coral Reefs section above, the more 
colonies a species has, the lower the proportion of colonies that are 
likely to be exposed to a particular threat at a particular time, and 
all individuals that are exposed will not have the same response.
    Notwithstanding the projections through 2100 that indicate 
increased severity over time of the three high importance threats, the 
combination of these biological and environmental characteristics 
indicates that the species possesses significant buffering capacity to 
avoid being in danger of extinction within the foreseeable future 
throughout its range. It is possible that M. tuberosa's extinction risk 
may increase in the future if global threats continue and worsen in 
severity, likely resulting in the continued decline of this species 
into the future. As the species experiences reduced abundance or range 
constriction of a certain magnitude, its ability to moderate exposure 
to threats will diminish. However, the species is not likely to become 
of such low abundance or so spatially fragmented as to be in danger of 
extinction due to depensatory processes, the potential effects of 
environmental stochasticity, or the potential for mortality from 
catastrophic events within the foreseeable future throughout its range. 
Therefore, M. tuberosa is not warranted for listing at this time under 
any of the listing factors.

Genus Seriatopora

Genus Introduction
    The family Pocilloporidae includes three genera: Pocillopora, 
Seriatopora, and Stylophora. Seriatopora contains six species, all 
occurring in the Indo-Pacific (Veron, 2000). Seriatopora species have 
branching colonies. The SRR and SIR provided no genus-level 
introductory information on Seriatopora.
Genus Susceptibility to Threats
    The SRR and SIR provided the following information on the threat 
susceptibilities of the genus Seriatopora. Species in the genus 
Seriatopora are highly susceptible to bleaching across regions, 
including Micronesia the GBR, and the western Indian Ocean. The genus 
Seriatopora is known to be susceptible to predation by snails and the 
crown-of-thorns seastar, Acanthaster planci. The genus Seriatopora has 
been heavily traded, primarily from Fiji and Indonesia (and 
occasionally the Philippines and Taiwan). Many records are at the genus 
level; trade was heavy in the mid-1980s (exceeding 134,000 pieces in 
1987). Seriatopora hystrix is the most heavily exploited species, 
although Seriatopora caliendrum is also exported.
    The public comments did not provide any new or supplemental 
information on the threat susceptibilities of the genus Seriatopora. We 
gathered supplemental information, which provided the following. There 
are several reports of high bleaching and mortality in Seriatopora 
species in response to warming events. In response to the 1998 warming 
event, Seriatopora colonies in Palau had high levels of bleaching with 
high mortality (Bruno et al., 2001). In response to the same warming 
event, over half of Seriatopora colonies in study sites within Kenyan 
marine protected areas were killed by mass bleaching (McClanahan et 
al., 2001). A large study of the bleaching responses of over 100 coral 
species on the GBR to the 1998 bleaching event included one Seriatopora 
species, Seriatopora hystrix. For this species, approximately 40 
percent of the observed colonies were bleached, resulting in S. hystrix 
being more affected than most of the Pocilloporidae and Acroporidae 
species in the study, and one of the 20 most affected species in the 
entire study (Done et al., 2003b).
    In response to a 2008 bleaching event in Papua New Guinea, two 
Pocilloporidae species (including S. hystrix) and 14 Acroporidae 
species were monitored, and each species' relative susceptibility to 
bleaching was evaluated in relationship to the other species in the 
study. Nine of the 16 species, including S. hystrix, had moderate 
susceptibility to bleaching, while five species were rated as severe or 
high susceptibilities, and two as low. Of the 139 S. hystrix colonies 
monitored in the study, 126 bleached (Bonin, 2012). In response to a 
2004 warming event in Mauritius, the genus Seriatopora was the most 
bleached of the 32 genera recorded (McClanahan et al., 2005b). In eight 
countries in the western Indian Ocean in 1998-2005, the Seriatopora 
genus had a bleaching index of 32, the fourth highest of the 45 genera 
recorded, and 75 percent of the highest value (McClanahan et al., 
2007a).
    McClanahan et al. (2007a) calculated a relative extinction risk 
score based on bleaching for genera of corals in the western Indian 
Ocean. The index of extinction risk was proportional to the degree of 
bleaching and inversely proportional to the abundance and number of 
reefs on which a taxon was found. The index of extinction risk for 
Seriatopora was the eighth highest out of 47 genera, with a score of 
0.46 based on a scale of zero to one, with one being the score of the 
highest extinction risk.
    With regard to disease, two reports from the GBR provide 
contrasting information regarding the susceptibilities of Seriatopora 
species to various coral diseases. One study found that Black Band 
Disease was nearly absent on colonies of Seriatopora species (Page and 
Willis, 2006). In contrast, colonies of Seriatopora species

[[Page 53980]]

had high susceptibility to Skeletal Eroding Band, with a prevalence of 
5.8 percent. Skeletal Eroding Band is the most prevalent disease on the 
GBR (Page and Willis, 2007). Seriatopora in Indonesia was reported to 
have no diseases (Haapkyla et al., 2007). There is no information 
available on the effects of any other threat for Seriatopora species.
Genus Conclusion
    Based on the information from the SRR, SIR, public comments, and 
supplemental information, we can make the following inferences about 
the susceptibilities of an unstudied Seriatopora species to ocean 
warming, disease, ocean acidification, sedimentation, nutrients, 
trophic effects of fishing, sea-level rise, predation, and collection 
and trade. The SRR rated ocean warming and disease as ``high'' 
importance to corals. These were rated as the three most important 
threats to reef-building corals overall. All studies on thermal stress 
in Seriatopora report high levels of bleaching in response to warming 
events. Thus, we conclude that Seriatopora likely has high 
susceptibility to ocean warming. Studies reported that one disease did 
not infect Seriatopora, but another did at high prevalence, and no 
diseases infected it in Indonesia. Thus, we conclude that Seriatopora 
has some susceptibility to disease. Although there is no other genus-
level or species-specific information on the susceptibilities of 
Seriatopora species to ocean acidification, the SRR rated it as 
``medium-high'' importance to corals. Thus, we conclude that an 
unstudied Seriatopora species has some susceptibility to ocean 
acidification.
    The SRR rated the trophic effects of fishing as ``medium'' 
importance, the fourth most important threat to corals overall. This 
threat was not addressed at the genus or species level in the SRR or 
SIR, because it is an ecosystem-level process. That is, removal of 
herbivorous fish from coral reef systems by fishing alters trophic 
interactions by reducing herbivory on algae, thereby providing a 
competitive advantage for space to algae over coral. Thus, the SRR did 
not discuss this threat in terms of coral taxa, as its effects are 
difficult to distinguish between coral genera and species. Therefore, 
we conclude that an unstudied Seriatopora species has some 
susceptibility to the trophic effects of fishing.
    Although there is no genus-level or species-specific information on 
the susceptibilities of Seriatopora species to sedimentation or 
nutrients, the SRR rated both threats as ``low-medium'' importance to 
corals. Thus, we conclude that an unstudied Seriatopora species has 
some susceptibility to these threats. Sea-level rise was also rated as 
``low-medium'' importance to corals. Increasing sea levels may increase 
land-based sources of pollution due to inundation, resulting in changes 
to coral community structure, thus an unstudied Seriatopora species is 
likely to have some susceptibility to sea-level rise. The SRR rated 
predation and ornamental trade (referred to in the proposed rule as 
Collection and Trade) as ``low'' importance to corals overall. 
Seriatopora is preyed on by both snails and crown-of-thorns starfish. 
Thus we conclude that Seriatopora has some susceptibility to predation. 
Seriatopora is heavily traded, thus we conclude that Seriatopora has 
some susceptibility to collection and trade.
    In conclusion, an unstudied Seriatopora species is likely to be 
highly susceptible to ocean warming, and to have some susceptibility to 
disease, ocean acidification, trophic effects of fishing, 
sedimentation, nutrients, sea-level rise, predation, and collection and 
trade.

Seriatopora aculeata

Introduction
    The SRR and SIR provided the following information on S. aculeata's 
morphology and taxonomy. Morphology was described as thick, short, 
tapered branches, usually in fused clumps. The taxonomy was described 
as somewhat uncertain, because genetic studies have not corresponded 
well with morphology for S. aculeata and other species of Seriatopora. 
Similar species, Seriatopora stellata and S. hystrix, can have similar 
branching structures in shallow, exposed reef flats.
    The public comments and information we gathered did not provide any 
new or supplemental information on morphology, and confirmed that there 
is a moderate level of taxonomic uncertainty for S. aculeata, and that 
there is a moderate level of species identification uncertainty for 
this species. Veron (Veron, 2014) states that S. aculeata is sometimes 
confused with S. stellata, but Veron (Veron, 2000; Veron, 2014) 
continues to consider it a valid species, and we conclude it can be 
identified by experts, and that the distribution and abundance 
information described below for this species is sufficiently reliable 
(Fenner, 2014b).
Spatial Information
    The SRR and SIR provided the following information on S. aculeata's 
distribution, habitat, and depth range. Seriatopora aculeata is 
distributed from Australia, Fiji, Indonesia, Japan, Papua New Guinea, 
and Madagascar to the Marshall Islands. The SRR and SIR described S. 
aculeata's habitat as shallow reef environments, and its depth range as 
three to 40 meters. The SIR reported it in Guam and the Northern 
Marianas.
    The public comments and information we gathered provided 
supplemental information on the distribution and habitat of S. 
aculeata. One public comment stated that in Guam, the few specimens of 
S. aculeata observed since 2004 were found in areas with high rates of 
sedimentation. Thus, based on all the available information, S. 
aculeata's habitat can be summarized as follows: The species occurs in 
a broad range of habitats on the reef slope and back-reef, including 
but not limited to upper reef slopes, mid-slope terraces, lower reef 
slopes, reef flats, and lagoons. Supplemental information provided the 
following. Veron (2014) provides an updated, much more detailed range 
map for this species than the maps used in the SRR. Veron reports that 
S. aculeata is confirmed in 19 of his 133 Indo-Pacific ecoregions, and 
strongly predicted to be found in an additional seven.
Demographic Information
    The SRR and SIR provided the following information on S. aculeata's 
abundance. Seriatopora aculeata has been reported as uncommon.
    The public comments did not provide any new or supplemental 
information on S. aculeata's abundance, but the supplemental 
information provided the following. Veron (2014) reports that S. 
aculeata occupied 10.3 percent of 2,984 dive sites sampled in 30 
ecoregions of the Indo-Pacific, and had a mean abundance rating of 1.70 
on a 1 to 5 rating scale at those sites in which it was found. Based on 
this semi-quantitative system, the species' abundance was characterized 
as ``common,'' and overall abundance was described as ``uncommon.'' 
Veron did not infer trends in abundance from these data. As described 
in the Indo-Pacific Species Determinations introduction above, based on 
results from Richards et al. (2008) and Veron (2014), the absolute 
abundance of this species is likely at least millions of colonies.
    Carpenter et al. (2008) extrapolated species abundance trend 
estimates from total live coral cover trends and habitat types. For S. 
aculeata, the overall decline in abundance (``Percent Population 
Reduction'') was estimated

[[Page 53981]]

at 37 percent, and the decline in abundance before the 1998 bleaching 
event (``Back-cast Percent Population Reduction'') was estimated at 14 
percent. However, as summarized above in the Inter-basin Comparison 
sub-section, live coral cover trends are highly variable both spatially 
and temporally, producing patterns on small scales that can be easily 
taken out of context, thus quantitative inferences to species-specific 
trends should be interpreted with caution. At the same time, an 
extensive body of literature documents broad declines in live coral 
cover and shifts to reef communities dominated by hardier coral species 
or algae over the past 50 to 100 years (Birkeland, 2004; Fenner, 2012; 
Pandolfi et al., 2003; Sale and Szmant, 2012). These changes have 
likely occurred, and are occurring, from a combination of global and 
local threats. Given that S. aculeata occurs in many areas affected by 
these broad changes, and that it has some susceptibility to both global 
and local threats, we conclude that it is likely to have declined in 
abundance over the past 50 to 100 years, but quantification is not 
possible based on the limited species-specific information.
Other Biological Information
    The SRR and SIR provided the following information on S. aculeata's 
life history. Little is known of S. aculeata's life history. The much 
more common species, S. hystrix, is a simultaneous hermaphrodite that 
reproduces sexually via brooded larvae. The public comments and 
information we gathered provided no additional biological information.
Susceptibility to Threats
    To describe S. aculeata's threats, the SRR and SIR provided genus-
level information for the effects on Seriatopora of o ocean warming, 
disease, acidification, sedimentation, nutrients, predation, and 
collection and trade. The SRR and SIR did not provide any species-
specific information on the effects of these threats on S. aculeata, 
except for a single export record from Indonesia for four pieces of the 
species in 2008. We interpreted the threat susceptibility and exposure 
information from the SRR and SIR in the proposed rule for S. aculeata's 
vulnerabilities as follows. High vulnerability to ocean warming; 
moderate vulnerability to disease, ocean acidification, trophic effects 
of reef fishing, nutrients, and predation; and low vulnerability to 
sedimentation, sea level rise, and collection and trade.
    Public comments provided some supplemental information on S. 
aculeata's threat susceptibilities. One comment stated that the depth 
range for S. aculeata on the reef slopes of Guam are coincident with 
those of the crown-of-thorns starfish, both of which are below 5 to 7 
meters depth, exposing S. aculeata to predation. Seriatopora aculeata 
has been rated as not moderately or highly susceptible to bleaching and 
disease, but this rating is not based on species-specific data 
(Carpenter et al. 2008). There is no supplemental species-specific 
information for the susceptibility of S. aculeata to any threat. Based 
on information provided in the Seriatopora genus description above, S. 
aculeata is likely to be highly susceptible to ocean warming, and is 
likely to have some susceptibility to disease, ocean acidification, 
trophic effects of fishing, sedimentation, nutrients, sea-level rise, 
predation, and collection and trade. The available information does not 
support more precise ratings of the susceptibilities of S. aculeata to 
the threats.
Regulatory Mechanisms.
    In the proposed rule, we did not provide any species-specific 
information on the regulatory mechanisms or conservation efforts for S. 
aculeata. Criticisms of our approach received during public comment led 
us to the following analysis to attempt to analyze regulatory 
mechanisms on a species basis. Records confirm that S. aculeata occurs 
in 19 Indo-Pacific ecoregions that encompass 10 countries' EEZs. The 10 
countries are Federated States of Micronesia, France (French Pacific 
Island Territories), Indonesia, Japan, Palau, Papua New Guinea, 
Philippines, Solomon Islands, Timor-Leste, and the United States (CNMI, 
Guam, PRIAs). The regulatory mechanisms available to S. aculeata, 
described first as a percentage of the above countries that utilize 
them to any degree, and second as the percentage of those countries 
whose regulatory mechanisms are limited in scope, are as follows: 
General coral protection (40 percent with none limited in scope), coral 
collection (70 percent with 20 percent limited in scope), pollution 
control (30 percent with 20 percent limited in scope), fishing 
regulations on reefs (100 percent with none limited in scope), and 
managing areas for protection and conservation (100 percent with none 
limited in scope). The most common regulatory mechanisms in place for 
S. aculeata are reef fishing regulations and area management for 
protection and conservation. Coral collection laws are also heavily 
utilized for the species. General coral protection and pollution 
control laws are less common regulatory mechanisms for the management 
of S. aculeata.
Vulnerability to Extinction
    As explained above in the Risk Analyses section, a species' 
vulnerability to extinction results from the combination of its spatial 
and demographic characteristics, threat susceptibilities, and 
consideration of the baseline environment and future projections of 
threats. The SRR stated that the primary factor that increases the 
potential extinction risk is its high bleaching susceptibility. The 
genus Seriatopora is heavily traded, but not often identified to 
species. Heavy use in the aquarium trade implies the potential for 
local extirpation for this usually uncommon species.
    Subsequent to the proposed rule, we received and gathered 
supplemental species- or genus-specific information, described above, 
that expands our knowledge regarding the species abundance, 
distribution, and threat susceptibilities. We developed our assessment 
of the species' vulnerability to extinction using all the available 
information. As explained in the Risk Analyses section, our assessment 
in this final rule emphasizes the ability of the species' spatial and 
demographic traits to moderate or exacerbate its vulnerability to 
extinction, as opposed to the approach we used in the proposed rule, 
which emphasized the species' susceptibility to threats.
    The following characteristics of S. aculeata, in conjunction with 
the information described in the Corals and Coral Reefs section, Coral 
Habitat sub-section, and Threats Evaluation section above, affect its 
vulnerability to extinction currently and over the foreseeable future. 
Its geographic distribution is limited to parts of the Coral Triangle 
and the western equatorial Pacific Ocean. Despite the large number of 
islands and environments that are included in the species' range, this 
range exacerbates vulnerability to extinction over the foreseeable 
future because it is mostly limited to an area projected to have the 
most rapid and severe impacts from climate change and localized human 
impacts for coral reefs over the 21st century. Its depth range of 40 
meters moderates vulnerability to extinction over the foreseeable 
future because deeper areas of its range will usually have lower 
irradiance than surface waters, and acidification is generally 
predicted to accelerate most in waters that are deeper and cooler than 
those in which the species occurs. The species

[[Page 53982]]

occurs in a broad range of habitats on the reef slope and back-reef, 
including but not limited to upper reef slopes, mid-slope terraces, 
lower reef slopes, reef flats, and lagoons. This moderates 
vulnerability to extinction over the foreseeable future because the 
species is not limited to one habitat type but occurs in numerous types 
of reef environments that will, on local and regional scales, 
experience highly variable thermal regimes and ocean chemistry at any 
given point in time. There is not enough information about its 
abundance to determine if it moderates or exacerbates extinction. It is 
common and has at least millions of colonies, but the great majority of 
the population is within an area expected to be severely impacted by 
threats over the foreseeable future. While depth distribution and 
habitat variability moderate vulnerability to extinction, the 
combination of its geographic distribution and high susceptibility to 
ocean warming are likely to be more influential to the status of this 
species over the foreseeable future, because of the projected severity 
of ocean warming throughout the species' range in the foreseeable 
future, and its high susceptibility to this threat.
Listing Determination
    In the proposed rule, using the determination tool formula 
approach, S. aculeata was proposed for listing as threatened because 
of: High vulnerability to ocean warming (ESA Factor E); moderate 
vulnerability to disease (C) and acidification (E); uncommon 
generalized range wide abundance (E); moderate overall distribution 
(based on moderate geographic distribution and moderate depth 
distribution (E); and inadequacy of existing regulatory mechanisms (D).
    In this final rule, we maintain the listing determination for S. 
aculeata as threatened. Based on the best available information 
provided above on S. aculeata's spatial structure, demography, threat 
susceptibilities, and management indicate that it is likely to become 
endangered throughout its range within the foreseeable future, and thus 
warrants listing as threatened at this time, because:
    (1) Seriatopora aculeata is highly susceptible to ocean warming 
(ESA Factor E), and susceptible to disease (C) ocean acidification (E), 
trophic effects of fishing (A), nutrients (A, E), and collection and 
trade (B). In addition, existing regulatory mechanisms to address 
global threats that contribute to extinction risk for this species are 
inadequate (D); and
    (2) Seriatopora aculeata's distribution is constrained to the Coral 
Triangle and western equatorial Pacific, which is projected to have the 
most rapid and severe impacts from climate change and localized human 
impacts for coral reefs over the 21st century, as described in the 
Threats Evaluation. Multiple ocean warming events have already occurred 
within the western equatorial Pacific that suggest future ocean warming 
events may be more severe than average in this part of the world. A 
range constrained to this particular geographic area that is likely to 
experience severe and increasing threats indicates that a high 
proportion of the population of this species is likely to be exposed to 
those threats over the foreseeable future.
    The combination of these characteristics and projections of future 
threats indicates that the species is likely to be in danger of 
extinction within the foreseeable future throughout its range and 
warrants listing as threatened at this time due to factors A, C, D, and 
E.
    The available information above on S. aculeata's spatial structure, 
demography, threat susceptibilities, and management also indicate that 
the species is not currently in danger of extinction and thus does not 
warrant listing as Endangered because:
    (1) While half of S. aculeatas' range is within the Coral Triangle 
which increases its extinction risk as described above, its habitat 
includes various shallow reef environments down to 40 meters. This 
moderates vulnerability to extinction currently because the species is 
not limited to one habitat type but occurs in numerous types of reef 
environments that will, at local and regional scales, experience highly 
variable thermal regimes and ocean chemistry at any given point in 
time, as described in more detail in the Coral Habitat sub-section and 
Threats Evaluation section. There is no evidence to suggest that the 
species is so spatially fragmented that depensatory processes, 
environmental stochasticity, or the potential for catastrophic events 
currently pose a high risk to the survival of the species;
    (2) Seriatopora aculeata occurs down to at least 40 m so its depth 
range will provide some refugia from threats because deeper areas of 
its range will usually have lower irradiance than surface water, and 
acidification is generally predicted to accelerate most in waters that 
are deeper and cooler than those in which the species occurs; and
    (3) Seriatopora aculeata's absolute abundance is at least millions 
of colonies, which allows for variation in the responses of individuals 
to threats to play a role in moderating vulnerability to extinction for 
the species to some degree, as described in more detail in the Corals 
and Coral Reefs section. There is no evidence of depensatory processes 
such as reproductive failure from low density of reproductive 
individuals and genetic processes such as inbreeding affecting this 
species. Thus, its absolute abundance indicates it is currently able to 
avoid high mortality from environmental stochasticity, and mortality of 
a high proportion of its population from catastrophic events.
    The combination of these characteristics indicates that the species 
does not exhibit the characteristics of one that is currently in danger 
of extinction, as described previously in the Risk Analyses section, 
and thus does not warrant listing as endangered at this time.
    Range-wide, a multitude of conservation efforts are already broadly 
employed that are likely benefiting S. aculeata. However, considering 
the global scale of the most important threats to the species, and the 
ineffectiveness of conservation efforts at addressing the root cause of 
global threats (i.e., GHG emissions), we do not believe that any 
current conservation efforts or conservation efforts planned in the 
future will result in affecting the species status to the point at 
which listing is not warranted.

Genus Acropora, Indo-Pacific

Genus Introduction
    The SRR and SIR provided an introduction to Indo-Pacific Acropora, 
covering geological history, taxonomy, life history, and threat 
susceptibilities of the genus as a whole. Acropora colonies are usually 
branching, bushy, or plate-like, rarely encrusting or submassive. 
Acropora is by far the largest genus of corals with over 150 species, 
and dominates many reefs, making Acropora the most important single 
genus of corals in the world. Almost all species of Acropora are in the 
Indo-Pacific.
Genus Susceptibility to Threats
    The SRR and SIR provided the following information on genus-level 
threat susceptibilities for Indo-Pacific Acropora. Acropora are widely 
reported to be more sensitive to bleaching in response to high 
temperatures than other coral genera. Some studies report branching 
species of Acropora to bleach more than table species, but other 
studies do not find this. Bleaching mortality in Acropora can be very 
severe. Larval connectivity and survival of partially-dead colonies are 
probably

[[Page 53983]]

important in population recovery. Bleaching of Acropora has been 
followed by disease outbreaks and by reduced fecundity for a year or 
two. Fertilization and larval stages of Acropora are particularly 
sensitive to high temperatures.
    Ocean acidification decreases the rate of calcification in 
Acropora. For one species of Acropora in the Caribbean, decreases in 
growth rates on reefs over decades has been attributed to 
acidification. Acidification negatively affects a variety of stages of 
reproduction in Acropora.
    Acropora are vulnerable to most of the diseases that infect coral, 
and are more commonly affected by acute and lethal diseases (``white 
diseases'' or tissue loss) than other corals. Such lethal diseases have 
been the major cause of the loss of most Acropora in the Caribbean. The 
reduction of coral populations by disease leads to negative synergisms, 
as it reduces Acropora reproductive output and can lead to recruitment 
failure, making population recovery very difficult.
    Acropora are preferred prey for most predators that prey on coral, 
including the crown-of-thorns starfish, a variety of snails including 
Drupella, butterflyfish, and fireworms. Individual territorial 
butterflyfish can take 400-700 bites per hour, and butterflyfish 
densities can be 50-70 per 1000m\2\, demonstrating possible intense 
predation on Acropora. Acropora have low carbon and protein content in 
their tissues so a low nutrient value, yet are still preferred prey. 
This suggests that instead of investing in chemical defenses against 
predation, Acropora invests its energy in rapid growth. However, when 
coral populations are greatly reduced, the predatory pressure is 
increased on colonies, and can exert a positive-feedback effect (Allee 
Effect or depensation) that makes populations unstable and can lead to 
collapse or lack of recovery.
    In general, Acropora species are relatively more susceptible to the 
effects of sedimentation than many other reef-building corals. Though 
certain growth forms (e.g., cylindrical branches) may be more effective 
at passive sediment rejection than others, Acropora are generally not 
adept at actively removing sediment. Acropora have also shown 
particular sensitivity to shading, an effect of turbid waters resulting 
from sedimentation. In addition, adult colonies of Acropora have 
reportedly shown impacts from sedimentation especially during 
reproduction.
    Acropora species are also relatively more susceptible to the 
effects of nutrients, especially with regard to reproduction and 
recruitment. Elevated nutrients have been shown to reduce fertilization 
success, survival, and settlement of Acropora larvae. Further, iron-
rich ``red'' soils typical of tropical islands, as well as other 
chemicals in run-off, interfere with synchronization of spawning among 
colonies, egg-sperm recognition and interactions, fertilization, and 
embryo development.
    Acropora species are heavily collected and widely traded 
internationally. Trade quotas and reports are typically listed only at 
the genus level, making any species-specific inferences with regard to 
this threat very difficult.
    The public comments did not provide any supplemental information on 
genus-level threat susceptibilities for Indo-Pacific Acropora. However, 
we gathered supplemental information, which provides the following 
genus-level information on threat susceptibilities of Indo-Pacific 
Acropora for ocean warming, disease, ocean acidification, and 
predation. With regard to susceptibility to ocean warming, Fisk and 
Done (1985) report bleaching patterns on a site on the Great Barrier 
Reef in 1982 to 1983. Most species of Acropora in shallow water had 
significant mortality, but Acropora hyacinthus did not. Mortality 
varied by species and site. Brown and Suharsono (1990) reported that 
the 1983 El Ni[ntilde]o caused a mass bleaching event in the Thousand 
Islands, Indonesia. The mass bleaching event killed all Acropora (22 
species) in the transects on the reef flats of two islands (Brown and 
Suharsono, 1990). Gleason (1993) reported that Acropora was the second 
most affected genus by bleaching (Montastraea was the most affected) in 
Moorea, French Polynesia in 1991, and that it had the greatest 
mortality. McClanahan et al. (2001) report that almost all Acropora in 
study sites in Kenya were killed by mass bleaching in 1998. Kayanne et 
al. (2002) reported that in 1998 in the Ryukyu Islands of Japan, 
branching Acropora was susceptible to bleaching and mortality was high. 
The branching species in this study were primarily A. formosa (= A. 
muricata) and also A. pulchra and A. palifera (= Isopora palifera). 
Hughes et al. (2003) reported that 11 Acropora species ranged from 0 to 
100 percent affected by bleaching in Raiatea, French Polynesia, in 
2002. Done et al. (2003b) reported that 46 Acropora species ranged from 
0 to 44 percent affected by bleaching on the Great Barrier Reef in 
2003.
    Based on a bleaching index scaled from 0 to 100 (with 0 as no 
bleaching and 100 as complete bleaching), McClanahan et al. (2004) 
reported that during mass bleaching in 1998, Acropora had a higher 
index in Kenya (80) than in Australia (40); temperatures were higher in 
Kenya. Acropora in Mauritius had an index of 39, the fifth highest of 
the 32 genera recorded, following a 2004 bleaching event (McClanahan et 
al., 2005a). Acropora had an index of 28.9 for eight countries in the 
western Indian Ocean in 1998-2005, which was fifth highest of the 45 
genera recorded (McClanahan et al., 2007a). The abundance of Acropora 
after 1998 in the western Indian Ocean decreased strongly in proportion 
to the number of degree heating weeks in 1998 (McClanahan et al., 
2007b). Based on a bleaching index scaled from 0 to 250 (with 0 as no 
bleaching and 250 as complete bleaching), Pandolfi et al. (2011) report 
that Acropora bleached heavily in Kenya and moderately in Australia in 
1998, with scores of 225 and 120, respectively. Acropora had a moderate 
percentage of bleaching on Howland and Baker islands in the U.S. 
Pacific in early 2010, with 28.7 percent bleached on Baker and 47.7 
percent on Howland. Acropora was the fifth most-bleached genus out of 
14 genera, and was 60 percent as bleached as the most bleached genus 
(Vargas-Angel et al., 2011).
    During a mass-bleaching event in Western Australia in 2010-2011, 
Acropora had the highest mortality with 100 percent mortality of 
colonies larger than 10 cm diameter in size, and Montipora the second 
highest mortality, while massive and encrusting corals, such as Porites 
and faviids, had much higher survival rates. Colonies less than 10 cm 
diameter were not killed (Depczynski et al., 2012). Acropora in the 
turbid waters off Okinawa, Japan, experienced sharp drops in 
populations following the 1998 and 2010 mass bleaching episodes (Hongo 
and Yamano, 2013). Sutthacheep et al. (2013) report that all colonies 
of one species of Acropora were completely bleached at Laem Set at 
Samui Island in the western Gulf of Thailand in 1998 and 80 percent of 
the colonies of the other reef-building coral species were as well. In 
2010, 80 percent colonies of one species were completely bleached and 
all colonies of the other species were partly bleached. After the 1998 
bleaching event, 72 percent of colonies had complete mortality, and 
after the 2010 event, all bleached colonies had complete mortality.
    Bleaching does not always result in mortality, thus it is important 
to consider bleaching-induced mortality and bleaching rates from a 
single event, as well as the recovery of a population over time to a 
bleaching event. In Kenya

[[Page 53984]]

in 1998, mortality in Acropora was sixth highest of the 18 genera, and 
55 percent of the genus with the most mortality (McClanahan, 2004). 
Three species of Acropora were long-term winners following mass 
bleaching events in Japan (decreasing from 3.4 percent cover to 0 
percent then increasing to 3.5 percent; decreasing from 0.2 percent to 
0 percent and then increasing to 3.2 percent; decreasing from 1.2 
percent cover to 0 percent and then increasing to 0.7 percent), and one 
species was neither a winner or a loser (van Woesik et al., 2011). 
Bridge et al. (2013a) report that Acropora mortality after bleaching 
was higher than for all corals as a whole. Total coral mortality at 0 
to 2 m depth was 70 percent, while it was 90 percent for Acropora, and 
at 3 to 4 m depth it was 20 percent for all corals but 60 percent for 
Acropora (Bridge et al., 2013a).
    Species or genera that readily bleach but recover quickly are 
relatively resilient to warming-induced bleaching. For example, the 
genus Acropora received a +1 resilience score based on trait and 
process scores assigned to the genus (van Woesik et al., 2012). Traits 
and processes were chosen which were thought to confer resilience to 
climate change. Resilience scores of 16 Indo-Pacific genera that were 
evaluated varied between +7 and -5. Scores below 0 were correlated with 
a high extinction probability (van Woesik et al., 2012). McClanahan et 
al. (2007a) calculated a relative extinction risk score based on 
bleaching for genera of corals in the western Indian Ocean. The index 
of extinction risk was proportional to the degree of bleaching and 
inversely proportional to the abundance and number of reefs on which a 
taxon was found. The index of extinction risk for Acropora was the 
ninth lowest out of 47 genera, with a score of 0.11 based on a scale of 
0 to 1, with 1 being the score of the highest extinction risk 
(McClanahan et al., 2007a).
    Diseases have been reported to be more common in Acropora than in 
other corals in some areas of the Indo-Pacific, such as the Northwest 
Hawaiian Islands (Aeby, 2006) and American Samoa (Fenner et al., 2008). 
However, in the Philippines, Porites was the dominant host with almost 
all disease observed in that genus, and only rarely observed on 
Acropora (Raymundo et al., 2005). In New Caledonia, Turbinaria had the 
highest disease prevalence of any genus with 2.5% infected, while 
Acropora was tied with Montipora for the least disease among the 12 
most common genera affected, with less than 0.1% infected (Tribollet et 
al., 2011). On the Great Barrier Reef, Pocilloporidae and Acroporidae 
have the highest prevalence of families, and diseases have been 
recorded on at least 23 species of Acropora (Willis et al., 2004). 
Black band disease on the Great Barrier Reef is concentrated in 
staghorn Acropora species with 76 diseased colonies counted in one 
study, and Acropora species with other colony morphologies (tables, 
bushy, corymbose, digitate, bottlebrush) had far fewer diseased 
colonies (Page and Willis, 2006). In American Samoa, French Frigate 
Shoals (Hawaii) and Johnston Atoll, two species of table Acropora (A. 
hyacinthus and A. cytherea) had larger numbers of colonies (13 each) 
with growth anomalies in transects than any of 10 other taxa, and much 
higher than one other table coral (A. clathrata, with one; Work et al., 
2008). In Indonesia, bushy Acropora had the highest prevalence (8%) of 
disease of any taxon (out of 35 taxa), while corymbose Acropora was the 
eighth highest taxon and second highest Acropora group with 0.5 percent 
disease, and all other Acropora groups (tabulate, bottlebrush, 
digitate, and staghorn) had 0 percent disease (Haapkyla et al., 2007).
    Ocean acidification can have a variety of effects on Indo-Pacific 
Acropora species. While increased CO2 does not appear to 
affect the survival of unidentified Acropora larvae, postsettlement 
skeletal growth of the polyps of unidentified Acropora species (Suwa et 
al., 2010) and A. digitifera (Inoue et al., 2011) are impaired. In 
addition, increased CO2 impairs the rate of zooxanthellae 
acquisition in the polyps of A. digitifera (Inoue et al., 2011) and A. 
millepora (Kaniewska et al., 2012). In Caribbean Acropora species, 
fertilization and settlement are impaired by increased CO2 
(Albright et al., 2010). Elevated CO2 also induces bleaching 
in Acropora, even more so than temperature increases (Anthony et al., 
2008). Carbon dioxide enrichment to 600 to 790 ppm enhanced maximum 
photosynthetic rates in A. formosa (Crawley et al., 2010), but elevated 
CO2 levels had no effect on photosynthesis or respiration in 
A. eurystoma (Schneider and Erez, 2006). A study of the effects of 
near-term ocean acidification and elevated seawater temperature on the 
physiology of A. aspera suggested that gene expression of key metabolic 
proteins is impacted by the synergistic effects of near term ocean 
acidification (i.e., the conditions expected to result from 50-90 ppm 
CO2 above current atmospheric levels) and ocean warming 
(Ogawa et al., 2013a). Physical factors may moderate impacts of 
acidification, as shown by a study of A. hyacinthus, which found that 
natural daily oscillations in CO2 may reduce the locally 
negative effects of increasing ocean acidification (Comeau et al., 
2014). Moderate increases in CO2 may enhance Acropora growth 
and calcification rates in some species, however, at higher 
CO2 levels, growth and calcification rates drop to zero. 
More consistently across species, elevated CO2 tends to 
decrease Acropora growth and calcification rates (Anthony et al., 2008; 
Chauvin et al., 2011; Purkis et al., 2011; Schneider and Erez, 2006; 
Suggett et al., 2013). Acropora species appear to be more susceptible 
to acidification than most other genera, as demonstrated by the lack of 
Acropora species in coral communities existing in naturally low pH 
waters (Fabricius et al., 2011).
    With regard to predation, De'ath and Moran (1998) reported that 
Acropora was the most preferred prey of crown-of-thorns starfish out of 
the 10 most common genera on 15 reefs in the Great Barrier Reef 
(preferred 14:1 over Porites, the least preferred genus). Pratchett 
(2001) reported that in a choice experiment, crown-of-thorns starfish 
always ate Acropora colonies before eating colonies of other genera. 
This was true of all four of the Acropora species tested. When a crown-
of-thorns starfish has finished eating preferred species, it moves to 
eating less preferred species, and thus in an outbreak, almost all 
species may be eaten (Pratchett et al., 2001). The snail Drupella 
rugosa preferred to eat Acropora pruinosa over Montipora informis, one 
agaricid and four faviid corals in laboratory tests in Hong Kong 
(Morton et al., 2002).
    The public comments did not provide any supplemental information on 
genus-level threat susceptibilities for Indo-Pacific Acropora. We 
gathered the supplemental information above, which provides genus-level 
information on threat susceptibilities of Indo-Pacific Acropora for 
ocean warming, disease, ocean acidification, and predation. We did not 
gather any supplemental information on the other threats (i.e., 
sedimentation, nutrients, trophic effects of fishing, sea-level rise, 
or collection and trade).
Genus Conclusion
    Based on the information from the SRR, SIR, public comments, and 
supplemental information, we make the following inferences regarding 
the susceptibilities of an unstudied Acropora species to ocean warming, 
disease, ocean acidification, predation, sedimentation, nutrients, 
trophic effects of fishing, sea-level rise, and collection

[[Page 53985]]

and trade. Nearly all the studies cited on thermal stress in Acropora 
report high levels of bleaching in response to warming events. Thus, it 
is possible to predict that an unstudied Acropora species is likely to 
be highly susceptible to warming-induced bleaching, as long as some 
considerations are kept in mind: (1) Despite high overall 
susceptibility within the genus to warming-induced bleaching, there can 
be high variability between species and habitats (Done et al., 2003b); 
(2) colonies that bleach do not necessarily die (in general, Acropora 
species have higher post-bleaching mortality than corals as a whole, 
but there is high variability in response throughout the genus); (3) 
recovery from bleaching provides the mechanism for acclimatization; and 
(4) while most Acropora species readily bleach in response to warming 
events, most also have the capacity to reestablish local populations 
relatively quickly through their rapid growth and asexual reproduction 
capacity.
    The studies cited above suggest that diseases are generally more 
common in Acropora than other coral genera, although there are numerous 
documented exceptions, depending on location. These studies also 
demonstrate high variability in disease susceptibility across Acropora 
species, depending on growth form, with wide divergence of disease 
susceptibilities among colony morphological groups under the same 
conditions. Thus, it is possible to predict that an unstudied Acropora 
species is likely to have some susceptibility to disease.
    The studies cited above on ocean acidification in Acropora report 
impacts on skeletal growth rates. Thus, it is possible to predict that 
an unstudied Acropora species is likely to have some susceptibility to 
ocean acidification in terms of impacts on skeletal growth. The studies 
cited above on predation in Acropora report that predators such as 
crown-of-thorns starfish and Drupella snails prefer to eat Acropora 
over other genera. Thus, it is possible to predict that an unstudied 
Acropora species is likely to have some susceptibility to predation. 
Most studies summarized in the SRR on the effects of land-based sources 
of pollution suggest that an unstudied Acropora species is likely to 
have some susceptibility to sedimentation and nutrient enrichment.
    The SRR rated the trophic effects of fishing as ``medium'' 
importance, and it was the fourth most important threat to corals 
overall. This threat was not addressed at the genus or species level in 
the SRR or SIR, because it is an ecosystem-level process. That is, 
removal of herbivorous fish from coral reef systems by fishing alters 
trophic interactions by reducing herbivory on algae, thereby providing 
a competitive advantage for space to algae over coral. Thus, the SRR 
did not discuss this threat in terms of coral taxa, as its effects are 
difficult to distinguish between coral genera and species. Accordingly, 
an unstudied Acropora species is likely to have some susceptibility to 
the trophic effects of fishing.
    The SRR rated sea-level rise as ``low-medium'' importance to corals 
overall. This threat was not addressed at the genus or species level in 
the SRR or SIR. Increasing sea levels may provide new coral habitats by 
submergence of hard substrates; however sea-level rise is also likely 
to increase land-based sources of pollution due to inundation, 
resulting in changes to coral community structure, most likely to 
sediment-tolerant assemblages and slower-growing species. Because 
Acropora are not generally sediment-tolerant and are faster growing 
species, an unstudied Acropora species is likely to have some 
susceptibility to sea-level rise.
    The SRR rated ornamental trade (referred to in the proposed rule as 
Collection and Trade) as ``low'' importance to corals overall, and this 
threat was addressed at both the genus and species levels in the SRR. 
Because Acropora species are some of the most popular coral species to 
collect and trade, an unstudied Acropora species is likely to have some 
susceptibility to collection and trade.
    In conclusion, an unstudied Acropora species is likely to be highly 
susceptible to ocean warming and to have some susceptibility to 
disease, acidification, sedimentation, nutrients, trophic effects of 
fishing, sea-level rise, predation, and collection and trade.

Acropora aculeus

Introduction
    The SRR and SIR provided the following information on A. aculeus' 
morphology and taxonomy. Morphology was described as small bushy 
colonies with flat tops, and taxonomy was described as having no 
taxonomic issues but being similar in appearance to A. latistella.
    The public comments and information we gathered provided 
information on the morphology or taxonomy of A. aculeus. One public 
comment stated that specimens collected in the Mariana Islands and 
identified by coral expert Richard H. Randall as A. aculeus appear to 
be different than colonies described as A. aculeus in references used 
in the SRR. Also, one public comment stated that specimens collected in 
American Samoa and identified by the American Samoa Department of 
Marine and Water Resources as A. jacquelineae appear to be A. aculeus, 
thereby illustrating the species identification uncertainties 
associated with this species. In addition, we gathered supplemental 
information, including Veron (2014), which states that this species is 
distinctive. Thus, while the public comments and supplemental 
information provided some information on the taxonomy of A. aculeus, we 
conclude it can be identified by experts, and that the distribution and 
abundance information described below for this species is sufficiently 
reliable (Fenner, 2014b).
 Spatial Information
    The SRR and SIR provided the following information on A. aculeus' 
distribution, habitat, and depth range: Acropora aculeus is distributed 
from East Africa to the Pitcairn Islands in the eastern Pacific. The 
SRR and SIR reported the species as having the 15th largest range of 
114 Acropora species in a large study. Its predominant habitat is 
shallow lagoons, and it is also found in other habitats protected from 
direct wave action on back-reefs and reef slopes, and its depth range 
is low tide to at least 20 m.
    The public comments did not provide any new or supplemental 
information on A. aculeus' distribution. We gathered supplemental 
information, including Veron (2014), which reports that this species is 
confirmed in 68 of his 133 Indo-Pacific ecoregions, and strongly 
predicted to be found in an additional 16. Wallace (1999b) reports its 
occurrence in 24 of her 29 Indo-Pacific areas, many of which are 
significantly larger than Veron's ecoregions. Richards (2009) 
calculated the geographic range of A. aculeus at over 100 million 
km\2\. The public comments and information we gathered provided nothing 
additional on A. aculeus' habitat and depth range.
Demographic Information
    The SRR and SIR provided the following information on A. aculeus' 
abundance. Acropora aculeus has been reported as generally common and 
locally abundant, especially in the central Indo-Pacific, and that it 
is particularly abundant in shallow lagoons and common in most habitats 
where it is protected from direct wave action.
    The public comments did not provide any new or supplemental 
information on A. aculeus' abundance. We gathered supplemental 
information, including

[[Page 53986]]

Richards (2009) and Richards et al. (2013b), which concluded that this 
species is globally widespread, locally widespread, and locally common. 
Based on these results, the authors concluded that A. aculeus is among 
the most abundant Acropora species, and also among those Acropora 
species that are most likely to persist in the future. They placed 12 
species in this category out of 85 species of Acropora. Veron (2014) 
reports that A. aculeus occupied 32.1 percent of 2,984 dive sites 
sampled in 30 ecoregions of the Indo-Pacific, and had a mean abundance 
rating of 1.55 on a 1 to 5 rating scale at those sites in which it was 
found. Based on this semi-quantitative system, the species' abundance 
was characterized as ``common.'' Overall abundance was described as 
``usually common in the central Indo-Pacific, uncommon elsewhere.'' 
Veron did not infer abundance trend results from these data. As 
described in the Indo-Pacific Species Determinations introduction 
above, based on results from Richards et al. (2008) and Veron (2014), 
the absolute abundance of this species is likely at least tens of 
millions of colonies.
    Carpenter et al. (2008) extrapolated species abundance trend 
estimates from total live coral cover trends and habitat types. For A. 
aculeus, the overall decline in abundance (``Percent Population 
Reduction'') was estimated at 37 percent, and the decline in abundance 
before the 1998 bleaching event (``Back-cast Percent Population 
Reduction'') was estimated at 15 percent. However, as summarized above 
in the Inter-basin Comparisons sub-section, live coral cover trends are 
highly variable both spatially and temporally, producing patterns on 
small scales that can be easily taken out of context. Thus quantitative 
inferences to species-specific trends should be interpreted with 
caution. At the same time, an extensive body of literature documents 
broad declines in live coral cover and shifts to reef communities 
dominated by hardier coral species or algae over the past 50 to 100 
years (Birkeland, 2004; Fenner, 2012; Pandolfi et al., 2003; Sale and 
Szmant, 2012). These changes have likely occurred and are occurring 
from a combination of global and local threats. Given that A. aculeus 
occurs in many areas affected by these broad changes, and that it is 
likely has some susceptibility to both global and local threats, we 
conclude that it is likely to have declined in abundance over the past 
50 to 100 years, but quantification is not possible based on the 
limited species-specific information.
Other Biological Information
    The SRR and SIR provided the following information on A. aculeus' 
life history. Acropora aculeus is a hermaphroditic spawner that is a 
participant in mass broadcast spawning in some localities. The public 
comments and information we gathered provided no additional biological 
information.
Susceptibility to Threats
    To describe A. aculeus' threat susceptibilities, the SRR and SIR 
provided genus-level information for the effects on Acropora of ocean 
warming, acidification, disease, predation, sedimentation, nutrients, 
and collection and trade. The SRR and SIR did not provide any other 
species-specific information on the effects of these threats on A. 
aculeus. We interpreted the threat susceptibility and exposure 
information from the SRR and SIR in the proposed rule for A. aculeus' 
vulnerabilities as follows: High vulnerability to ocean warming, 
moderate vulnerabilities to disease, acidification, trophic effects of 
fishing, nutrient over-enrichment, and predation, and low 
vulnerabilities to sedimentation, sea-level rise, and collection and 
trade.
    Public comments provided some supplemental information on A. 
aculeus' threat susceptibilities. One comment stated that A. aculeus is 
more susceptible to predation than indicated in the proposed rule 
because of the overlap in the depth ranges of this species with crown 
of thorns starfish. In addition, we gathered the following species-
specific and genus-level supplemental information on this species' 
threat susceptibilities. Acropora aculeus has been rated as moderately 
or highly susceptible to bleaching, but this rating is not based on 
species-specific data (Carpenter et al., 2008). Done et al. (2003b) 
report 20 percent of A. aculeus colonies were affected by bleaching on 
the GBR in 2002, and the species ranked 31st in proportion of coral 
colonies on the GBR that were bleached and killed out of 52 studied 
Acropora species. That is, 30 of the 52 species bleached more than A. 
aculeus, and 21 bleached less. Bonin (2012) reported that A. aculeus 
had a ``high'' susceptibility to bleaching in Kimbe Bay, Papua New 
Guinea on a scale of ``severe,'' ``high,'' ``moderate,'' and 
``lowest.'' Acropora aculeus was fourth highest out of 16 species, with 
50 percent of colonies either severely bleached or dead. The most 
severely affected species had 74 percent of colonies either severely 
bleached or dead (Bonin, 2012).
    Acropora aculeus has been rated as moderately or highly susceptible 
to disease, but this rating is not based on species-specific data 
(Carpenter et al., 2008). Page and Willis (2007) reported that Skeletal 
Eroding Band has been found in A. aculeus. Skeletal Eroding Band is the 
most prevalent disease on the Great Barrier Reef. They also reported 
that corymbose Acropora had moderate susceptibility to Skeletal Eroding 
Band in the Great Barrier Reef, with a prevalence of 2.4 percent (Page 
and Willis, 2007). No other species-specific information is available 
for the susceptibility of A. aculeus to any other threat.
    Based on information from other Acropora species provided in the 
genus description above, A. aculeus may be susceptible to the effects 
of ocean acidification on skeletal growth. Genus-level information also 
suggests that A. aculeus is susceptible to trophic effects of fishing, 
sedimentation, nutrients, predation, sea-level rise, and collection and 
trade. Thus, based on the available species-specific and genus 
information summarized above, A. aculeus is likely highly susceptible 
to ocean warming, and also likely has some susceptibilities to disease, 
acidification, trophic effects of fishing, sedimentation, nutrients, 
predation, sea-level rise, and collection and trade. The available 
information does not support more precise ratings of the 
susceptibilities of A. aculeus to the threats.
Regulatory Mechanisms
    In the proposed rule we did not provide any species-specific 
information on the regulatory mechanisms or conservation efforts for A. 
aculeus. Public comments were critical of that approach, and we 
therefore attempt to analyze regulatory mechanisms and conservation 
efforts on a species basis, where possible, in this final rule. Records 
confirm that A. aculeus occurs in 68 Indo-Pacific ecoregions that 
encompass 39 countries' EEZs. The 39 countries are Australia, 
Bangladesh, Brunei, China, Comoros Islands, Federated States of 
Micronesia, Fiji, France (French Pacific Island Territories), India 
(including Andaman and Nicobar Islands), Indonesia, Japan, Kenya, 
Kiribati, Madagascar, Malaysia, Maldives, Marshall Islands, Mauritius, 
Mozambique, Myanmar, New Zealand (Tokelau), Niue, Palau, Papua New 
Guinea, Philippines, Samoa, Seychelles, Solomon Islands, South Africa, 
Sri Lanka, Taiwan, Tanzania, Thailand, Tonga, Tuvalu, United Kingdom 
(British Indian Ocean Territory and Pitcairn Islands), United States 
(CNMI, Guam, American Samoa, PRIAs), Vanuatu, and

[[Page 53987]]

Vietnam. The regulatory mechanisms relevant to A. aculeus, described 
first as the percentage of the above countries that utilize them to any 
degree, and second as the percentages of those countries whose 
regulatory mechanisms may be limited in scope, are as follows: General 
coral protection (28 percent with 8 percent limited in scope), coral 
collection (56 percent with 31 percent limited in scope), pollution 
control (38 percent with 10 percent limited in scope), fishing 
regulations on reefs (95 percent with 26 percent limited in scope), and 
managing areas for protection and conservation (97 percent with 8 
percent limited in scope). The most common regulatory mechanisms in 
place for A. aculeus are reef fishing regulations and area management 
for protection and conservation. Coral collection laws are also 
somewhat common for the species, but 31 percent of coral collection 
laws are limited in scope and may not provide substantial protection. 
General coral protection and pollution control laws are much less 
common regulatory mechanisms for the management of A. aculeus.
Vulnerability to Extinction
    As explained above in the Risk Analyses section, a species' 
vulnerability to extinction results from the combination of its spatial 
and demographic characteristics, threat susceptibilities, and 
consideration of the baseline environment and future projections of 
threats. The SRR stated that the high bleaching rate of the Acropora 
genus is the primary known threat of extinction for A. aculeus. It 
listed factors that reduce A. aculeus' threat of extinction including 
its geographic range, depth range, abundance, and variable habitats.
    Subsequent to the proposed rule, we received and gathered 
supplemental species- or genus-specific information, described above, 
that expands our knowledge regarding the species abundance, 
distribution, and threat susceptibilities. We developed our assessment 
of the species' vulnerability to extinction using all the available 
information. As explained in the Risk Analyses section, our assessment 
in this final rule emphasizes the ability of the species' spatial and 
demographic traits to moderate or exacerbate its vulnerability to 
extinction, as opposed to the approach we used in the proposed rule, 
which emphasized the species' susceptibility to threats.
    The following characteristics of A. aculeus, in conjunction with 
the information described in the Corals and Coral Reefs section, Coral 
Habitat sub-section, and Threats Evaluation section above, affect its 
vulnerability to extinction currently and over the foreseeable future. 
Its geographic distribution includes most of the coral reef ecoregions 
in the Indian Ocean and western and central Pacific Ocean. Its 
geographic distribution moderates vulnerability to extinction because 
some areas within its range are projected to have less than average 
warming and acidification over the foreseeable future, including the 
western Indian Ocean, the central Pacific, and other areas, so portions 
of the population in these areas will be less exposed to severe 
conditions. Its depth range is from low tide to at least 20 meters. 
This moderates vulnerability to extinction over the foreseeable future 
because deeper areas of its range will usually have lower irradiance 
than surface waters, and acidification is generally predicted to 
accelerate most in waters that are deeper and cooler than those in 
which the species occurs. Its predominant habitat is shallow lagoons, 
and it is found in other habitats protected from direct wave action on 
back-reefs and reef slopes. This moderates vulnerability to extinction 
over the foreseeable future because the species is not limited to one 
habitat type but occurs in numerous types of reef environments that 
will, on local and regional scales, experience highly variable thermal 
regimes and ocean chemistry at any given point in time. Its absolute 
abundance of at least tens of millions of colonies, combined with 
spatial variability in ocean warming and acidification across the 
species range, moderates vulnerability to extinction because the 
increasingly severe conditions expected in the foreseeable future will 
be non-uniform and therefore will likely be a large number of colonies 
that are either not exposed or do not negatively respond to a threat at 
any given point in time.
Listing Determination
    In the proposed rule using the determination tool formula approach, 
A. aculeus was proposed for listing as threatened because of: High 
vulnerability to ocean warming (ESA Factor E); moderate vulnerability 
to disease (C) and acidification (E); common generalized range wide 
abundance (E); wide overall distribution (based on wide geographic 
distribution and moderate depth distribution (E); and inadequacy of 
existing regulatory mechanisms (D).
    In this final rule, we changed the listing determination for A. 
aculeus from threatened to not warranted. We made this determination 
based on a more species-specific and holistic assessment of whether 
this species meets the definition of either a threatened or endangered 
coral largely in response to public comments, including more 
appropriate consideration of the buffering capacity of this species' 
spatial and demographic traits to lessen its vulnerability to threats. 
Thus, based on the best available information above on A. aculeus' 
spatial structure, demography, threat susceptibilities, and management, 
none of the five ESA listing factors, alone or in combination, are 
causing this species to be likely to become endangered throughout its 
range within the foreseeable future, and thus it is not warranted for 
listing at this time, because:
    (1) Acropora aculeus' distribution across the Indian Ocean and most 
of the Pacific Ocean is spread over a very large area. While some areas 
within its range are projected to be affected by warming and 
acidification, other areas are projected to have less than average 
warming and acidification, including the western Indian Ocean, the 
central Pacific, and other areas. This distribution and the 
heterogeneous habitats it occupies reduce exposure to any given threat 
event or adverse condition that does not occur uniformly throughout the 
species range. As explained above in the Threats Evaluation section, we 
have not identified any threat that is expected to occur uniformly 
throughout the species range within the foreseeable future);
    (2) Acropora aculeus' total absolute abundance is at least tens of 
millions of colonies, providing buffering capacity in the form of 
absolute numbers of colonies and variation in susceptibility between 
individual colonies. As discussed in the Corals and Coral Reefs section 
above, the more colonies a species has, the lower the proportion of 
colonies that are likely to be exposed to a particular threat at a 
particular time, and all individuals that are exposed will not have the 
same response; and
    (3) It is a broadcast spawner and fast grower, enhancing recovery 
potential from mortality events as described in the Corals and Coral 
Reefs section above.
    Notwithstanding the projections through 2100 that indicate 
increased severity over time of the three high importance threats, the 
combination of these biological and environmental characteristics 
indicates that the species possesses sufficient buffering capacity to 
avoid being in danger of extinction within the foreseeable future 
throughout its range. It is possible that this species' extinction risk 
may increase in the future if global threats continue and

[[Page 53988]]

worsen in severity and the species' exposure to the threats increases 
throughout its range. Should the species experience reduced abundance 
or range constriction of a certain magnitude, the ability of these 
characteristics to moderate exposure to threats will diminish. However, 
the species is not likely to become of such low abundance or so 
spatially fragmented as to be in danger of extinction due to 
depensatory processes, the potential effects of environmental 
stochasticity, or the potential for mortality from catastrophic events 
within the foreseeable future throughout its range. Therefore, A. 
aculeus is not warranted for listing at this time under any of the 
listing factors.

Acropora acuminata

Introduction
    The SRR and SIR provided the following information on A. 
acuminata's morphology and taxonomy. Morphology was described as 
typically forming a tabular base of fused horizontal branches that turn 
upward and taper to points, and the taxonomy was described as having no 
taxonomic issues, but colonies turn black when dried.
    The public comments and information we gathered provided 
information on the morphology or taxonomy of A. acuminata. One public 
comment letter stated that specimens of A. acuminata in the Mariana 
Islands may be a different species or a distinct sub-species, based on 
colony morphology. We gathered supplemental information, including 
Veron (2014), which states that this species is distinctive. While the 
public comments and supplemental information provided some information 
on the morphology and taxonomy of A. acuminata, it is sufficiently 
distinctive to be identified by experts, thus we conclude that the 
distribution and abundance information described below for this species 
is sufficiently reliable (Fenner, 2014b).
 Spatial Information
    The SRR and SIR provided the following information on A. 
acuminata's distribution, habitat, and depth range. Acropora 
acuminata's distribution is from the Red Sea to the Pitcairn Islands in 
the eastern Pacific, covering 110 million km\2\, the 5th largest range 
of 114 Acropora species in a large study. In general, its habitat is 
upper reef slopes and mid-slope terraces and shelves in turbid or clear 
water at 15-20 m of depth. In Guam, its habitat is deeper reef flat 
areas and channel slopes.
    The public comments and information we gathered provided 
information on the distribution and habitat of A. acuminata. One public 
comment letter stated that A. acuminata in the Mariana Islands appears 
to be restricted to reef flats and upper reef slopes in protected to 
semi-protected areas. Thus, based on all the available information, A. 
acuminata's habitat can be summarized as follows: Its predominant 
habitat is upper reef slopes and mid-slope terraces and shelves in 
turbid or clear water, and it also occurs in back-reef habitats 
including reef flats and channels. Its depth range is approximately two 
to 20 m depth. We gathered supplemental information, including Veron 
(2014), which reports that A. acuminata is confirmed in 60 of his 133 
Indo-Pacific ecoregions and is strongly predicted to be found in an 
additional 12. Wallace (1999b) reports its occurrence in 23 of her 29 
Indo-Pacific areas, many of which are significantly larger than Veron's 
ecoregions.
Demographic Information
    The SRR and SIR provided the following information on A. 
acuminata's abundance. Acropora acuminata has been reported to 
occasionally live in extensive clumps with dimensions of several 
meters, and it can be very common in the center of its range (e.g., 
Indonesia), but it can be uncommon in the outer parts of its range. The 
public comments and information we gathered provided information on the 
abundance of A. acuminata. A public comment letter stated that A. 
acuminata in the Mariana Islands is uncommon to rare. We gathered 
supplemental information, including Richards (2009) and Richards et al. 
(2013b), which conclude from their data that this species is globally 
widespread, locally restricted, and locally rare, and thus in the 
second rarest category of Acropora with the predicted consequence of 
persistence. They placed 39 species in this category out of 85 species 
of Acropora. Veron (2014) reports that A. acuminata occupied 4.7 
percent of 2,984 dive sites sampled in 30 ecoregions of the Indo-
Pacific, and had a mean abundance rating of 1.21 on a 1 to 5 rating 
scale at those sites in which it was found. Based on this semi-
quantitative system, the species' abundance was characterized as 
``uncommon.'' Overall abundance was described as ``sometimes common.'' 
Veron did not infer trends in abundance from these data. As described 
in the Indo-Pacific Species Determinations introduction above, based on 
results from Richards et al. (2008) and Veron (2014), the absolute 
abundance of this species is likely at least tens of millions of 
colonies.
    Carpenter et al. (2008) extrapolated species abundance trend 
estimates from total live coral cover trends and habitat types. For A. 
acuminata, the overall decline in abundance (``Percent Population 
Reduction'') was estimated at 35 percent, and the decline in abundance 
before the 1998 bleaching event (``Back-cast Percent Population 
Reduction'') was estimated at 14 percent. However, as summarized above 
in the Inter-basin Comparison sub-section, live coral cover trends are 
highly variable both spatially and temporally, producing patterns on 
small scales that can be easily taken out of context. Thus, 
quantitative inferences to species-specific trends should be 
interpreted with caution. At the same time, an extensive body of 
literature documents broad declines in live coral cover and shifts to 
reef communities dominated by hardier coral species or algae over the 
past 50 to 100 years (Birkeland, 2004; Fenner, 2012; Pandolfi et al., 
2003; Sale and Szmant, 2012). These changes have likely occurred, and 
are occurring, from a combination of global and local threats. Given 
that A. acuminata occurs in many areas affected by these broad changes, 
and that it has some susceptibility to both global and local threats, 
we conclude that it is likely to have declined in abundance over the 
past 50 to 100 years, but a precise quantification is not possible 
based on the limited species-specific information.
Other Biological Information
    The SRR and SIR provided the following information on A. 
acuminata's life history. Like most of its congeners, A. acuminata is a 
broadcast spawner. However, some degree of reproductive isolation 
probably occurs in some locations because the species does not spawn 
synchronously with the majority of its congeners. The public comments 
and information we gathered provided no additional biological 
information.
Susceptibility to Threats
    To describe A. acuminata's threat susceptibilities, the SRR and SIR 
provided genus-level information for the effects on Acropora of ocean 
warming, disease, acidification, sedimentation, nutrients, predation, 
and collection and trade. The SRR and SIR also stated that Acropora 
acuminata is the only Acropora known to not be preferred as prey by the 
crown-of-thorns starfish, thus susceptibility to predation appears to 
be low. The SRR and SIR did not

[[Page 53989]]

provide any other species-specific i