Federal Register, Volume 76 Issue 147 (Monday, August 1, 2011)

[Federal Register Volume 76, Number 147 (Monday, August 1, 2011)]
[Proposed Rules]
[Pages 46083-46147]
From the Federal Register Online via the Government Printing Office [www.gpo.gov]
[FR Doc No: 2011-18582]

[[Page 46083]]

Vol. 76

Monday,

No. 147

August 1, 2011

Part III

Environmental Protection Agency

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40 CFR Part 50

Secondary National Ambient Air Quality Standards for Oxides of
Nitrogen and Sulfur; Proposed Rule

Federal Register / Vol. 76 , No. 147 / Monday, August 1, 2011 /
Proposed Rules

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ENVIRONMENTAL PROTECTION AGENCY

40 CFR Part 50

[EPA-HQ-OAR-2007-1145; FRL-9441-2]
RIN 2060-AO72

Secondary National Ambient Air Quality Standards for Oxides of
Nitrogen and Sulfur

AGENCY: Environmental Protection Agency (EPA).

ACTION: Proposed rule.

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SUMMARY: This proposed rule is being issued as required by a consent
decree governing the schedule for completion of this review of the air
quality criteria and the secondary national ambient air quality
standards (NAAQS) for oxides of nitrogen and oxides of sulfur. Based on
its review, EPA proposes to retain the current nitrogen dioxide
(NO2) and sulfur dioxide (SO2) secondary
standards to provide requisite protection for the direct effects on
vegetation resulting from exposure to gaseous oxides of nitrogen and
sulfur in the ambient air. Additionally, with regard to protection from
the deposition of oxides of nitrogen and sulfur to sensitive aquatic
and terrestrial ecosystems, including acidification and nutrient
enrichment effects, EPA is proposing to add secondary standards
identical to the NO2 and SO2 primary 1-hour
standards and not set a new multi-pollutant secondary standard in this
review. The proposed 1-hour secondary NO2 standard would be
set at a level of 100 ppb and the proposed 1-hour secondary
SO2 standard would be set at 75 ppb. In addition, EPA has
decided to undertake a field pilot program to gather and analyze
additional relevant data so as to enhance the Agency's understanding of
the degree of protectiveness that a new multi-pollutant approach,
defined in terms of an aquatic acidification index (AAI), would afford
and to support development of an appropriate monitoring network for
such a standard. The EPA solicits comment on the framework of such a
standard and on the design of the field pilot program. The EPA will
sign a notice of final rulemaking for this review no later than March
20, 2012.

DATES: Written comments on this proposed rule must be received by
September 30, 2011.
Public Hearings: The EPA intends to hold a public hearing around
the end of August to early September and will announce in a separate
Federal Register notice the date, time, and address of the public
hearing on this proposed rule.

ADDRESSES: Submit your comments, identified by Docket ID No. EPA-HQ-
OAR-2007-1145, by one of the following methods:
http://www.regulations.gov: Follow the on-line
instructions for submitting comments.
E-mail: a-and-r-Docket@epa.gov.
Fax: 202-566-1741.
Mail: Docket No. EPA-HQ-OAR-2007-1145, Environmental
Protection Agency, Mail code 6102T, 1200 Pennsylvania Ave., NW.,
Washington, DC 20460. Please include a total of two copies.
Hand Delivery: Docket No. EPA-HQ-OAR-2007-1145,
Environmental Protection Agency, EPA West, Room 3334, 1301 Constitution
Ave., NW., Washington, DC. Such deliveries are only accepted during the
Docket's normal hours of operation, and special arrangements should be
made for deliveries of boxed information.
Instructions: Direct your comments to Docket ID No. EPA-HQ-OAR-
2007-1145. The EPA's policy is that all comments received will be
included in the public docket without change and may be made available
online at http://www.regulations.gov, including any personal
information provided, unless the comment includes information claimed
to be Confidential Business Information (CBI) or other information
whose disclosure is restricted by statute. Do not submit information
that you consider to be CBI or otherwise protected through http://www.regulations.gov or e-mail. The http://www.regulations.gov Web site
is an ``anonymous access'' system, which means EPA will not know your
identity or contact information unless you provide it in the body of
your comment. If you send an e-mail comment directly to EPA without
going through http://www.regulations.gov, your e-mail address will be
automatically captured and included as part of the comment that is
placed in the public docket and made available on the Internet. If you
submit an electronic comment, EPA recommends that you include your name
and other contact information in the body of your comment and with any
disk or CD-ROM you submit. If EPA cannot read your comment due to
technical difficulties and cannot contact you for clarification, EPA
may not be able to consider your comment. Electronic files should avoid
the use of special characters, any form of encryption, and be free of
any defects or viruses. For additional information about EPA's public
docket, visit the EPA Docket Center homepage at http://www.epa.gov/epahome/dockets.htm.
Docket: All documents in the docket are listed in the http://www.regulations.gov index. Although listed in the index, some
information is not publicly available, e.g., CBI or other information
whose disclosure is restricted by statute. Certain other material, such
as copyrighted material, will be publicly available only in hard copy.
Publicly available docket materials are available either electronically
in http://www.regulations.gov or in hard copy at the Air and Radiation
Docket and Information Center, EPA/DC, EPA West, Room 3334, 1301
Constitution Ave., NW., Washington, DC. The Public Reading Room is open
from 8:30 a.m. to 4:30 p.m., Monday through Friday, excluding legal
holidays. The telephone number for the Public Reading Room is (202)
566-1744 and the telephone number for the Air and Radiation Docket and
Information Center is (202) 566-1742.

FOR FURTHER INFORMATION CONTACT: Dr. Richard Scheffe, Office of Air
Quality Planning and Standards, U.S. Environmental Protection Agency,
Mail code C304-02, Research Triangle Park, NC 27711; telephone: 919-
541-4650; fax: 919-541-2357; e-mail: scheffe.rich@epa.gov.

SUPPLEMENTARY INFORMATION:

General Information

What should I consider as I prepare my comments for EPA?

1. Submitting CBI. Do not submit this information to EPA through
http://www.regulations.gov or e-mail. Clearly mark the part or all of
the information that you claim to be CBI. For CBI information in a disk
or CD ROM that you mail to EPA, mark the outside of the disk or CD ROM
as CBI and then identify electronically within the disk or CD ROM the
specific information that is claimed as CBI. In addition to one
complete version of the comment that includes information claimed as
CBI, a copy of the comment that does not contain the information
claimed as CBI must be submitted for inclusion in the public docket.
Information so marked will not be disclosed except in accordance with
procedures set forth in 40 CFR part 2.
2. Tips for Preparing Your Comments. When submitting comments,
remember to:
Identify the rulemaking by docket number and other
identifying information (subject heading, Federal Register date and
page number).

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Follow directions--The Agency may ask you to respond to
specific questions or organize comments by referencing a Code of
Federal Regulations (CFR) part or section number.
Explain why you agree or disagree, suggest alternatives,
and substitute language for your requested changes.
Describe any assumptions and provide any technical
information and/or data that you used.
If you estimate potential costs or burdens, explain how
you arrived at your estimate in sufficient detail to allow for it to be
reproduced.
Provide specific examples to illustrate your concerns, and
suggest alternatives.
Explain your views as clearly as possible.
Make sure to submit your comments by the comment period
deadline identified.

Availability of Related Information

A number of documents relevant to this rulemaking are available on
EPA web sites. The Integrated Science Assessment for Oxides of Nitrogen
and Sulfur--Ecological Criteria: Final Report (ISA) is available on
EPAs National Center for Environmental Assessment Web site. To obtain
this document, go to http://www.epa.gov/ncea, and click on Air Quality
then click on Oxides of Nitrogen and Sulfur. The Policy Assessment
(PA), Risk and Exposure Assessment (REA), and other related technical
documents are available on EPA's Office of Air Quality Planning and
Standards (OAQPS) Technology Transfer Network (TTN) web site. The PA is
available at http://www.epa.gov/ttn/naaqs/standards/no2so2sec/cr_pa.html, and the exposure and risk assessments and other related
technical documents are available at http://www.epa.gov/ttn/naaqs/standards/no2so2sec/cr_rea.html. These and other related documents are
also available for inspection and copying in the EPA docket identified
above.

Table of Contents

The following topics are discussed in this preamble:

I. Background
A. Legislative Requirements
B. History of Reviews of NAAQS for Nitrogen Oxides and Sulfur
Oxides
1. NAAQS for Oxides of Nitrogen
2. NAAQS for Oxides of Sulfur
C. History of Related Assessments and Agency Actions
D. History of the Current Review
E. Scope of the Current Review
II. Rationale for Proposed Decision on the Adequacy of the Current
Secondary Standards
A. Ecological Effects
1. Effects Associated with Gas-Phase Oxides of Nitrogen and
Sulfur
a. Nature of ecosystem responses to gas-phase nitrogen and
sulfur
b. Magnitude of ecosystem response to gas-phase nitrogen and
sulfur
2. Acidification Effects Associated with Deposition of Oxides of
Nitrogen and Sulfur
a. Nature of Acidification-related Ecosystem Responses
i. Aquatic Ecosystems
ii. Terrestrial Ecosystems
iii. Ecosystem Sensitivity
b. Magnitude of Acidification-Related Ecosystem Responses
i. Aquatic Acidification
ii. Terrestrial Acidification
c. Key Uncertainties Associated With Acidification
i. Aquatic Acidification
ii. Terrestrial Acidification
3. Nutrient Enrichment Effects Associated With Deposition of
Oxides of Nitrogen
a. Nature of Nutrient Enrichment-Related Ecosystem Responses
i. Aquatic Ecosystems
ii. Terrestrial Ecosystems
iii. Ecosystem Eensitivity to Nutrient Enrichment
b. Magnitude of Nutrient Enrichment-Related Ecosystem Responses
i. Aquatic Ecosystems
ii. Terrestrial Ecosystems
c. Key Uncertainties Associated With Nutrient Enrichment
i. Aquatic Ecosystems
ii. Terrestrial Ecosystems
4. Other Ecological Effects
B. Risk and Exposure Assessment
1. Overview of Risk and Exposure Assessment
2. Key Findings
a. Air Quality Analyses
b. Deposition-Related Aquatic Acidification
c. Deposition-Related Terrestrial Acidification
d. Deposition-Related Aquatic Nutrient Enrichment
e. Deposition-Related Terrestrial Nutrient Enrichment
f. Additional Effects
3. Conclusions on Effects
C. Adversity of Effects to Public Welfare
1. Ecosystem Services
2. Effects on Ecosystem Services
a. Aquatic Acidification
b. Terrestrial Acidification
c. Nutrient Enrichment
3. Summary
D. Adequacy of the Current Standards
1. Adequacy of the Current Standards for Direct Effects
2. Appropriateness and Adequacy of the Current Standards for
Deposition-Related Effects
a. Appropriateness
b. Adequacy of Protection
i. Aquatic Acidification
ii. Terrestrial Acidification
iii. Terrestrial Nutrient Enrichment
iv. Aquatic Nutrient Enrichment
v. Other Effects
3. CASAC Views
4. Administrator's Proposed Conclusions Concerning Adequacy of
Current Standard
III. Rationale for Proposed Decision on Alternative Multi-Pollutant
Approach to Secondary Standards for Aquatic Acidification
A. Ambient Air Indicators
1. Oxides of Sulfur
2. Oxides of Nitrogen
B. Form
1. Ecological Indicator
2. Linking ANC to Deposition
3. Linking Deposition to Ambient Air Indicators
4. Aquatic Acidification Index
5. Spatial Aggregation
a. Ecoregion Sensitivity
b. Representative Ecoregion-Specific Factors
i. Factor F1
(a) Acid-Sensitive Ecoregions
(b) Non-Acid Sensitive Ecoregions
ii. Factor F2
iii. Factors F3 and F4
c. Factors in Data-limited Ecoregions
d. Application to Hawaii, Alaska, and the U.S. Territories
6. Summary of the AAI Form
C. Averaging Time
D. Level
1. Association Between pH Levels and Target ANC Levels
2. ANC Levels Related to Effects on Aquatic Ecosystems
3. Consideration of Episodic Acidity
4. Consideration of Ecosystem Response Time
5. Prior Examples of Target ANC Levels
6. Consideration of Public Welfare Benefits
7. Summary of Alternative Levels
E. Combined Alternative Levels and Forms
F. Characterization of Uncertainties
1. Overview of Uncertainty
2. Uncertainties Associated with Data Gaps
3. Uncertainties in Modeled Processes
4. Applying Knowledge of Uncertainties
G. CASAC Advice
H. Administrator's Proposed Conclusions
IV. Field Pilot Program and Ambient Monitoring
A. Field Pilot Program
1. Objectives
2. Overview of Field Pilot Program
3. Complementary Measurements
4. Complementary Areas of Research Implementation Challenges
5. Final Monitoring Plan Development and Stakeholder
Participation
B. Evaluation of Monitoring Methods
1. Potential FRMs for SO2 and p-SO4
2. Potential FRM for NOy
V. Statutory and Executive Order Reviews
A. Executive Order 12866: Regulatory Planning and Review
B. Paperwork Reduction Act
C. Regulatory Flexibility Act
D. Unfunded Mandates Reform Act
E. Executive Order 13132: Federalism
F. Executive Order 13175: Consultation and Coordination With
Indian Tribal Governments
G. Executive Order 13045: Protection of Children From
Environmental Health and Safety Risks

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H. Executive Order 13211: Actions That Significantly Affect
Energy Supply, Distribution, or Use
I. National Technology Transfer and Advancement Act
J. Executive Order 12898: Federal Actions To Address
Environmental Justice in Minority Populations and Low-Income
Populations References

I. Background

A. Legislative Requirements

Two sections of the Clean Air Act (CAA) govern the establishment
and revision of the NAAQS. Section 108 (42 U.S.C. section 7408) directs
the Administrator to identify and list certain air pollutants and then
to issue air quality criteria for those pollutants. The Administrator
is to list those air pollutants that in her ``judgment, cause or
contribute to air pollution which may reasonably be anticipated to
endanger public health or welfare;'' ``the presence of which in the
ambient air results from numerous or diverse mobile or stationary
sources;'' and ``for which * * * [the Administrator] plans to issue air
quality criteria * * *'' Air quality criteria are intended to
``accurately reflect the latest scientific knowledge useful in
indicating the kind and extent of all identifiable effects on public
health or welfare which may be expected from the presence of [a]
pollutant in the ambient air * * *'' 42 U.S.C. 7408(b). Section 109 (42
U.S.C. 7409) directs the Administrator to propose and promulgate
``primary'' and ``secondary'' NAAQS for pollutants for which air
quality criteria are issued. Section 109(b)(1) defines a primary
standard as one ``the attainment and maintenance of which in the
judgment of the Administrator, based on such criteria and allowing an
adequate margin of safety, are requisite to protect the public
health.'' \1\ A secondary standard, as defined in section 109(b)(2),
must ``specify a level of air quality the attainment and maintenance of
which, in the judgment of the Administrator, based on such criteria, is
requisite to protect the public welfare from any known or anticipated
adverse effects associated with the presence of [the] pollutant in the
ambient air.'' Welfare effects as defined in section 302(h) (42 U.S.C.
7602(h)) include, but are not limited to, ``effects on soils, water,
crops, vegetation, man-made materials, animals, wildlife, weather,
visibility and climate, damage to and deterioration of property, and
hazards to transportation, as well as effects on economic values and on
personal comfort and well-being.''
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\1\ The legislative history of section 109 indicates that a
primary standard is to be set at ``the maximum permissible ambient
air level * * * which will protect the health of any [sensitive]
group of the population,'' and that for this purpose ``reference
should be made to a representative sample of persons comprising the
sensitive group rather than to a single person in such a group.'' S.
Rep. No. 91-1196, 91st Cong., 2d Sess. 10 (1970).
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In setting standards that are ``requisite'' to protect public
health and welfare, as provided in section 109(b), EPA's task is to
establish standards that are neither more nor less stringent than
necessary for these purposes. In so doing, EPA may not consider the
costs of implementing the standards. See generally, Whitman v. American
Trucking Associations, 531 U.S. 457, 465-472, 475-76 (2001). Likewise,
``[a]ttainability and technological feasibility are not relevant
considerations in the promulgation of national ambient air quality
standards.'' American Petroleum Institute v. Costle, 665 F. 2d at 1185.
Section 109(d)(1) requires that ``not later than December 31, 1980, and
at 5-year intervals thereafter, the Administrator shall complete a
thorough review of the criteria published under section 108 and the
national ambient air quality standards * * * and shall make such
revisions in such criteria and standards and promulgate such new
standards as may be appropriate * * * .'' Section 109(d)(2) requires
that an independent scientific review committee ``shall complete a
review of the criteria * * * and the national primary and secondary
ambient air quality standards * * * and shall recommend to the
Administrator any new * * * standards and revisions of existing
criteria and standards as may be appropriate * * * .'' Since the early
1980's, this independent review function has been performed by the
Clean Air Scientific Advisory Committee (CASAC).

B. History of Reviews of NAAQS for Nitrogen Oxides and Sulfur Oxides

1. NAAQS for Oxides of Nitrogen
After reviewing the relevant science on the public health and
welfare effects associated with oxides of nitrogen, EPA promulgated
identical primary and secondary NAAQS for NO2 in April 1971.
These standards were set at a level of 0.053 parts per million (ppm) as
an annual average (36 FR 8186). In 1982, EPA published Air Quality
Criteria Document for Oxides of Nitrogen (US EPA, 1982), which updated
the scientific criteria upon which the initial standards were based. In
February 1984 EPA proposed to retain these standards (49 FR 6866).
After taking into account public comments, EPA published the final
decision to retain these standards in June 1985 (50 FR 25532).
The EPA began the most recent previous review of the oxides of
nitrogen secondary standards in 1987. In November 1991, EPA released an
updated draft air quality criteria document (AQCD) for CASAC and public
review and comment (56 FR 59285), which provided a comprehensive
assessment of the available scientific and technical information on
health and welfare effects associated with NO2 and other
oxides of nitrogen. The CASAC reviewed the draft document at a meeting
held on July 1, 1993 and concluded in a closure letter to the
Administrator that the document ``provides a scientifically balanced
and defensible summary of current knowledge of the effects of this
pollutant and provides an adequate basis for EPA to make a decision as
to the appropriate NAAQS for NO2'' (Wolff, 1993). The AQCD
for Oxides of Nitrogen was then finalized (US EPA, 1995a). The EPA's
OAQPS also prepared a Staff Paper that summarized and integrated the
key studies and scientific evidence contained in the revised AQCD for
oxides of nitrogen and identified the critical elements to be
considered in the review of the NO2 NAAQS. The CASAC
reviewed two drafts of the Staff Paper and concluded in a closure
letter to the Administrator that the document provided a
``scientifically adequate basis for regulatory decisions on nitrogen
dioxide'' (Wolff, 1995).
In October 1995, the Administrator announced her proposed decision
not to revise either the primary or secondary NAAQS for NO2
(60 FR 52874; October 11, 1995). A year later, the Administrator made a
final determination not to revise the NAAQS for NO2 after
careful evaluation of the comments received on the proposal (61 FR
52852; October 8, 1996). While the primary NO2 standard was
revised in January 2010 by supplementing the existing annual standard
with the establishment of a new 1-hour standard, set at a level of 100
ppb (75 FR 6474), the secondary NAAQS for NO2 remains 0.053
ppm (100 micrograms per cubic meter [[mu]g/m3] of air), annual
arithmetic average, calculated as the arithmetic mean of the 1-hour
NO2 concentrations.
2. The NAAQS for Oxides of Sulfur
The EPA promulgated primary and secondary NAAQS for SO2
in April 1971 (36 FR 8186). The secondary standards included a standard
set at 0.02 ppm, annual arithmetic mean, and a 3-hour average standard
set at 0.5 ppm, not to be exceeded more than once per year. These
secondary standards

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were established solely on the basis of evidence of adverse effects on
vegetation. In 1973, revisions made to Chapter 5 (``Effects of Sulfur
Oxide in the Atmosphere on Vegetation'') of the AQCD for Sulfur Oxides
(US EPA, 1973) indicated that it could not properly be concluded that
the vegetation injury reported resulted from the average SO2
exposure over the growing season, rather than from short-term peak
concentrations. Therefore, EPA proposed (38 FR 11355) and then
finalized (38 FR 25678) a revocation of the annual mean secondary
standard. At that time, EPA was aware that then-current concentrations
of oxides of sulfur in the ambient air had other public welfare
effects, including effects on materials, visibility, soils, and water.
However, the available data were considered insufficient to establish a
quantitative relationship between specific ambient concentrations of
oxides of sulfur and such public welfare effects (38 FR 25679).
In 1979, EPA announced that it was revising the AQCD for oxides of
sulfur concurrently with that for particulate matter (PM) and would
produce a combined PM and oxides of sulfur criteria document. Following
its review of a draft revised criteria document in August 1980, CASAC
concluded that acid deposition was a topic of extreme scientific
complexity because of the difficulty in establishing firm quantitative
relationships among (1) Emissions of relevant pollutants (e.g.,
SO2 and oxides of nitrogen), (2) formation of acidic wet and
dry deposition products, and (3) effects on terrestrial and aquatic
ecosystems. The CASAC also noted that acid deposition involves, at a
minimum, several different criteria pollutants: Oxides of sulfur,
oxides of nitrogen, and the fine particulate fraction of suspended
particles. The CASAC felt that any document on this subject should
address both wet and dry deposition, since dry deposition was believed
to account for a substantial portion of the total acid deposition
problem.
For these reasons, CASAC recommended that a separate, comprehensive
document on acid deposition be prepared prior to any consideration of
using the NAAQS as a regulatory mechanism for the control of acid
deposition. The CASAC also suggested that a discussion of acid
deposition be included in the AQCDs for oxides of nitrogen and PM and
oxides of sulfur. Following CASAC closure on the AQCD for oxides of
sulfur in December 1981, EPA's OAQPS published a Staff Paper in
November 1982, although the paper did not directly assess the issue of
acid deposition. Instead, EPA subsequently prepared the following
documents to address acid deposition: The Acidic Deposition Phenomenon
and Its Effects: Critical Assessment Review Papers, Volumes I and II
(US EPA, 1984a, b) and The Acidic Deposition Phenomenon and Its
Effects: Critical Assessment Document (US EPA, 1985) (53 FR 14935-
14936). These documents, though they were not considered criteria
documents and did not undergo CASAC review, represented the most
comprehensive summary of scientific information relevant to acid
deposition completed by EPA at that point.
In April 1988 (53 FR 14926), EPA proposed not to revise the
existing primary and secondary standards for SO2. This
proposed decision with regard to the secondary SO2 NAAQS was
due to the Administrator's conclusions that: (1) Based upon the then-
current scientific understanding of the acid deposition problem, it
would be premature and unwise to prescribe any regulatory control
program at that time; and (2) when the fundamental scientific
uncertainties had been decreased through ongoing research efforts, EPA
would draft and support an appropriate set of control measures.
Although EPA revised the primary SO2 standard in June 2010
by establishing a new 1-hour standard at a level of 75 ppb and revoking
the existing 24-hour and annual standards (75 FR 35520), no further
decisions on the secondary SO2 standard have been published.

C. History of Related Assessments and Agency Actions

In 1980, the Congress created the National Acid Precipitation
Assessment Program (NAPAP) in response to growing concern about acidic
deposition. The NAPAP was given a broad 10-year mandate to examine the
causes and effects of acidic deposition and to explore alternative
control options to alleviate acidic deposition and its effects. During
the course of the program, the NAPAP issued a series of publicly
available interim reports prior to the completion of a final report in
1990 (NAPAP, 1990).
In spite of the complexities and significant remaining
uncertainties associated with the acid deposition problem, it soon
became clear that a program to address acid deposition was needed. The
Clean Air Act Amendments of 1990 included numerous separate provisions
related to the acid deposition problem. The primary and most important
of the provisions, the amendments to Title IV of the Act, established
the Acid Rain Program to reduce emissions of SO2 by 10
million tons and emissions of nitrogen oxides by 2 million tons from
1980 emission levels in order to achieve reductions over broad
geographic regions. In this provision, Congress included a statement of
findings that led them to take action, concluding that (1) The presence
of acid compounds and their precursors in the atmosphere and in
deposition from the atmosphere represents a threat to natural
resources, ecosystems, materials, visibility, and public health; (2)
the problem of acid deposition is of national and international
significance; and (3) current and future generations of Americans will
be adversely affected by delaying measures to remedy the problem.
Second, Congress authorized the continuation of the NAPAP in order
to assure that the research and monitoring efforts already undertaken
would continue to be coordinated and would provide the basis for an
impartial assessment of the effectiveness of the Title IV program.
Third, Congress considered that further action might be necessary
in the long term to address any problems remaining after implementation
of the Title IV program and, reserving judgment on the form that action
could take, included Section 404 of the 1990 Amendments (Clean Air Act
Amendments of 1990, Pub. L. 101-549, Sec. 404) requiring EPA to
conduct a study on the feasibility and effectiveness of an acid
deposition standard or standards to protect ``sensitive and critically
sensitive aquatic and terrestrial resources.'' At the conclusion of the
study, EPA was to submit a report to Congress. Five years later, EPA
submitted its report, entitled Acid Deposition Standard Feasibility
Study: Report to Congress (US EPA, 1995b) in fulfillment of this
requirement. That report concluded that establishing acid deposition
standards for sulfur and nitrogen deposition may at some point in the
future be technically feasible, although appropriate deposition loads
for these acidifying chemicals could not be defined with reasonable
certainty at that time.
Fourth, the 1990 Amendments also added new language to sections of
the CAA pertaining to the scope and application of the secondary NAAQS
designed to protect the public welfare. Specifically, the definition of
``effects on welfare'' in Section 302(h) was expanded to state that the
welfare effects include effects ``* * * whether caused by
transformation, conversion, or combination with other air pollutants.''

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In 1999, seven Northeastern states cited this amended language in
Section 302(h) in a petition asking EPA to use its authority under the
NAAQS program to promulgate secondary NAAQS for the criteria pollutants
associated with the formation of acid rain. The petition stated that
this language ``clearly references the transformation of pollutants
resulting in the inevitable formation of sulfate and nitrate aerosols
and/or their ultimate environmental impacts as wet and dry deposition,
clearly signaling Congressional intent that the welfare damage
occasioned by sulfur and nitrogen oxides be addressed through the
secondary standard provisions of Section 109 of the Act.'' The petition
further stated that ``recent federal studies, including the NAPAP
Biennial Report to Congress: An Integrated Assessment, document the
continued and increasing damage being inflicted by acid deposition to
the lakes and forests of New York, New England and other parts of our
nation, demonstrating that the Title IV program had proven
insufficient.'' The petition also listed other adverse welfare effects
associated with the transformation of these criteria pollutants,
including impaired visibility, eutrophication of coastal estuaries,
global warming, and tropospheric ozone and stratospheric ozone
depletion.
In a related matter, the Office of the Secretary of the U.S.
Department of Interior (DOI) requested in 2000 that EPA initiate a
rulemaking proceeding to enhance the air quality in national parks and
wilderness areas in order to protect resources and values that are
being adversely affected by air pollution. Included among the effects
of concern identified in the request were the acidification of streams,
surface waters, and/or soils; eutrophication of coastal waters;
visibility impairment; and foliar injury from ozone.
In a Federal Register notice in 2001 (65 FR 48699), EPA announced
receipt of these requests and asked for comment on the issues raised in
them. The EPA stated that it would consider any relevant comments and
information submitted, along with the information provided by the
petitioners and DOI, before making any decision concerning a response
to these requests for rulemaking.
The 2005 NAPAP report states that ``* * * scientific studies
indicate that the emission reductions achieved by Title IV are not
sufficient to allow recovery of acid-sensitive ecosystems. Estimates
from the literature of the scope of additional emission reductions that
are necessary in order to protect acid-sensitive ecosystems range from
approximately 40-80% beyond full implementation of Title IV. * * *''
The results of the modeling presented in this Report to Congress
indicate that broader recovery is not predicted without additional
emission reductions (NAPAP, 2005).
Given the state of the science as described in the ISA, REA, and in
other recent reports, such as the NAPAP reports noted above, EPA has
decided, in the context of evaluating the adequacy of the current
NO2 and SO2 secondary standards in this review,
to revisit the question of the appropriateness of setting secondary
NAAQS to address remaining known or anticipated adverse public welfare
effects resulting from the acidic and nutrient deposition of these
criteria pollutants.

D. History of the Current Review

The EPA initiated this current review in December 2005 with a call
for information (70 FR 73236) for the development of a revised ISA. An
Integrated Review Plan (IRP) was developed to provide the framework and
schedule as well as the scope of the review and to identify policy-
relevant questions to be addressed in the components of the review. The
IRP was released in 2007 (US EPA, 2007) for CASAC and public review.
The EPA held a workshop in July 2007 on the ISA to obtain broad input
from the relevant scientific communities. This workshop helped to
inform the preparation of the first draft ISA, which was released for
CASAC and public review in December 2007; a CASAC meeting was held on
April 2-3, 2008 to review the first draft ISA. A second draft ISA was
released for CASAC and public review in August 2008, and was discussed
at a CASAC meeting held on October 1-2, 2008. The final ISA (US EPA,
2008) was released in December 2008.
Based on the science presented in the ISA, EPA developed the REA to
further assess the national impact of the effects documented in the
ISA. The Draft Scope and Methods Plan for Risk/Exposure Assessment:
Secondary NAAQS Review for Oxides of Nitrogen and Oxides of Sulfur
outlining the scope and design of the future REA was prepared for CASAC
consultation and public review in March 2008. A first draft REA was
presented to CASAC and the public for review in August 2008 and a
second draft was presented for review in June 2009. The final REA (US
EPA, 2009) was released in September 2009. A first draft PA was
released in March 2010 and reviewed by CASAC on April 1-2, 2010. In a
June 22, 2010 letter to the Administrator, CASAC provided advice and
recommendations to the Agency concerning the first draft PA (Russell
and Samet, 2010a). A second draft PA was released to CASAC and the
public in September 2010 and reviewed by CASAC on October 6-7, 2010.
The CASAC provided advice and recommendations to the Agency regarding
the second draft PA in a December 9, 2010 letter (Russell and Samet
2010b). The CASAC and public comments on the second draft PA were
considered by EPA staff in developing a final PA (US EPA, 2011). CASAC
requested an additional meeting to provide additional advice to the
Administrator based on the final PA on February 15-16, 2011. On January
14, 2011, EPA released a version of the final PA prior to final
document production, to provide sufficient time for CASAC review of the
document in advance of this meeting. The final PA, incorporating final
reference checks and document formatting, was released in February
2011. In a May 17, 2011 letter (Russell and Samet, 2011a), CASAC
offered additional advice and recommendations to the Administrator with
regard to the review of the secondary NAAQS for oxides of nitrogen and
oxides of sulfur.
In 2005, the Center for Biological Diversity and four other
plaintiffs filed a complaint alleging that EPA had failed to complete
the current review within the period provided by statute.\2\ The
schedule for completion of this review is governed by a consent decree
resolving that lawsuit and the subsequent extension agreed to by the
parties. The schedule presented in the original consent decree that
governs this review, entered by the court on November 19, 2007, was
revised on October 22, 2009 to allow for a 17-month extension of the
schedule. The current decree provides that EPA sign for publication
notices of proposed and final rulemaking concerning its review of the
oxides of nitrogen and oxides of sulfur NAAQS no later than July 12,
2011 and March 20, 2012, respectively.
---------------------------------------------------------------------------

\2\ Center for Biological Diversity, et al. v. Johnson, No. 05-
1814 (D.D.C.)
---------------------------------------------------------------------------

This action presents the Administrator's proposed decisions on the
review of the current secondary oxides of nitrogen and oxides of sulfur
standards. Throughout this preamble a number of conclusions, findings,
and determinations proposed by the Administrator are noted. While they
identify the reasoning that supports this proposal, they are only
proposals and are not intended to be final or conclusive in nature. The
EPA invites general, specific, and/or technical

[[Page 46089]]

comments on all issues involved with this proposal, including all such
proposed judgments, conclusions, findings, and determinations.

E. Scope of the Current Review

In conducting this periodic review of the secondary NAAQS for
oxides of nitrogen and oxides of sulfur, as discussed in the IRP and
REA, EPA decided to assess the scientific information, associated
risks, and standards relevant to protecting the public welfare from
adverse effects associated jointly with oxides of nitrogen and sulfur.
Although EPA has historically adopted separate secondary standards for
oxides of nitrogen and oxides of sulfur, EPA is conducting a joint
review of these standards because oxides of nitrogen and sulfur, and
their associated transformation products are linked from an atmospheric
chemistry perspective, as well as from an environmental effects
perspective. The National Research Council (NRC) has recommended that
EPA consider multiple pollutants, as appropriate, in forming the
scientific basis for the NAAQS (NRC, 2004). As discussed in the ISA and
REA, there is a strong basis for considering these pollutants together,
building upon EPA's past recognition of the interactions of these
pollutants and on the growing body of scientific information that is
now available related to these interactions and associated ecological
effects.
In defining the scope of this review, it must be considered that
EPA has set secondary standards for two other criteria pollutants
related to oxides of nitrogen and sulfur: Ozone and particulate matter
(PM). Oxides of nitrogen are precursors to the formation of ozone in
the atmosphere, and under certain conditions, can combine with
atmospheric ammonia to form ammonium nitrate, a component of fine PM.
Oxides of sulfur are precursors to the formation of particulate
sulfate, which is a significant component of fine PM in many parts of
the U.S. There are a number of welfare effects directly associated with
ozone and fine PM, including ozone-related damage to vegetation and PM-
related visibility impairment. Protection against those effects is
provided by the ozone and fine PM secondary standards. This review
focuses on evaluation of the protection provided by secondary standards
for oxides of nitrogen and sulfur for two general types of effects: (1)
Direct effects on vegetation associated with exposure to gaseous oxides
of nitrogen and sulfur in the ambient air, which are the effects that
the current NO2 and SO2 secondary standards
protect against; and (2) effects associated with the deposition of
oxides of nitrogen and sulfur to sensitive aquatic and terrestrial
ecosystems, including deposition in the form of particulate nitrate and
particulate sulfate.
The ISA focuses on the ecological effects associated with
deposition of ambient oxides of nitrogen and sulfur to natural
sensitive ecosystems, as distinguished from commercially managed
forests and agricultural lands. This focus reflects the fact that the
majority of the scientific evidence regarding acidification and
nutrient enrichment is based on studies in unmanaged ecosystems. Non-
managed terrestrial ecosystems tend to have a higher fraction of
nitrogen deposition resulting from atmospheric nitrogen (US EPA, 2008,
section 3.3.2.5). In addition, the ISA notes that agricultural and
commercial forest lands are routinely fertilized with amounts of
nitrogen that exceed air pollutant inputs even in the most polluted
areas (US EPA, 2008, section 3.3.9). This review recognizes that the
effects of nitrogen deposition in managed areas are viewed differently
from a public welfare perspective than are the effects of nitrogen
deposition in natural, unmanaged ecosystems, largely due to the more
homogeneous, controlled nature of species composition and development
in managed ecosystems and the potential for benefits of increased
productivity in those ecosystems.
In focusing on natural sensitive ecosystems, the PA primarily
considers the effects of ambient oxides of nitrogen and sulfur via
deposition on multiple ecological receptors. The ISA highlights effects
including those associated with acidification and nitrogen nutrient
enrichment. With a focus on these deposition-related effects, EPA's
objective is to develop a framework for oxides of nitrogen and sulfur
standards that incorporates ecologically relevant factors and that
recognizes the interactions between the two pollutants as they deposit
to sensitive ecosystems. The overarching policy objective is to develop
a secondary standard(s) based on the ecological criteria described in
the ISA and the results of the assessments in the REA, and consistent
with the requirement of the CAA to set secondary standards that are
requisite to protect the public welfare from any known or anticipated
adverse effects associated with the presence of these air pollutants in
the ambient air. Consistent with the CAA, this policy objective
includes consideration of ``variable factors * * * which of themselves
or in combination with other factors may alter the effects on public
welfare'' of the criteria air pollutants included in this review.
In addition, we have chosen to focus on the effects of ambient
oxides of nitrogen and sulfur on ecological impacts on sensitive
aquatic ecosystems associated with acidifying deposition of nitrogen
and sulfur, which is a transformation product of ambient oxides of
nitrogen and sulfur. Based on the information in the ISA, the
assessments presented in the REA, and advice from CASAC on earlier
drafts of this PA (Russell and Samet, 2010a, 2010b), and as discussed
in detail in the PA, we have the greatest confidence in the causal
linkages between oxides of nitrogen and sulfur and aquatic
acidification effects relative to other deposition-related effects,
including terrestrial acidification and aquatic and terrestrial
nutrient enrichment.

II. Rationale for Proposed Decision on the Adequacy of the Current
Secondary Standards

Decisions on retaining or revising the current secondary standards
for oxides of nitrogen and sulfur are largely public welfare policy
judgments based on the Administrator's informed assessment of what
constitutes requisite protection against adverse effects to public
welfare. A public welfare policy decision should draw upon scientific
information and analyses about welfare effects, exposure and risks, as
well as judgments about the appropriate response to the range of
uncertainties that are inherent in the scientific evidence and
analyses. The ultimate determination as to what level of damage to
ecosystems and the services provided by those ecosystems is adverse to
public welfare is not wholly a scientific question, although it is
informed by scientific studies linking ecosystem damage to losses in
ecosystem services, and information on the value of those losses of
ecosystem services. In reaching such decisions, the Administrator seeks
to establish standards that are neither more nor less stringent than
necessary for this purpose.
This section presents the rationale for the Administrator's
proposed conclusions with regard to the adequacy of protection and
ecological relevance of the current secondary standards for oxides of
nitrogen and sulfur. As discussed more fully below, this rationale
considered the latest scientific information on ecological effects
associated with the presence of oxides of nitrogen and oxides of sulfur
in the ambient air. This rationale also takes into account: (1) Staff
assessments of the most policy-relevant information in the ISA and
staff analyses of air quality,

[[Page 46090]]

exposure, and ecological risks, presented more fully in the REA and in
the PA, upon which staff conclusions on revisions to the secondary
oxides of nitrogen and oxides of sulfur standards are based; (2) CASAC
advice and recommendations, as reflected in discussions of drafts of
the ISA, REA, and PA at public meetings, in separate written comments,
and in CASAC's letters to the Administrator; and (3) public comments
received during the development of these documents, either in
connection with CASAC meetings or separately.
In developing this rationale, EPA has drawn upon an integrative
synthesis of the entire body of evidence, published through early 2008,
on ecological effects associated with the deposition of oxides of
nitrogen and oxides of sulfur in the ambient air (US EPA, 2008). As
discussed below in section II.A, this body of evidence addresses a
broad range of ecological endpoints associated with ambient levels of
oxides of nitrogen and oxides of sulfur. In considering this evidence,
EPA focuses on those ecological endpoints, such as aquatic
acidification, for which the ISA judges associations with oxides of
nitrogen and oxides of sulfur to be causal, likely causal, or for which
the evidence is suggestive that oxides of nitrogen and/or sulfur
contribute to the reported effects. The categories of causality
determinations have been developed in the ISA (US EPA, 2008) and are
discussed in Section 1.6 of the ISA.
Crucial to this review is the development of a form for an
ecologically relevant standard that reflects both the geographically
variable and deposition-dependent nature of the effects. The
atmospheric levels of oxides of nitrogen and sulfur that afford a
particular level of ecosystem protection are those levels that result
in an amount of deposition that is less than the amount of deposition
that a given ecosystem can accept without defined levels of
degradation.
Drawing from the framework developed in the REA, the framework we
used to structure an ecologically meaningful secondary standard in the
PA and to further develop the indicator, form, level, and averaging
time of such a standard in section III of this proposal is depicted
below and highlights the three key linkages that need to be considered
in developing an ecologically relevant standard.
[GRAPHIC] [TIFF OMITTED] TP01AU11.023

The following discussion relies heavily on chapters 2 and 3 of the
PA. The PA includes staff's evaluation of the policy implications of
the scientific assessment of the evidence presented and assessed in the
ISA and the results of quantitative assessments based on that
information presented and assessed in the REA. Taken together, this
information informs staff conclusions and the development of policy
options in the PA for consideration in addressing public and welfare
effects associated with the presence of oxides of nitrogen and oxides
of sulfur in the ambient air. Of particular note, chapter 2 of the PA
presents information not repeated here that characterizes emissions,
air quality, deposition and water quality. It includes discussions of
the sources of nitrogen and sulfur in the atmosphere as well as current
ambient air quality monitoring networks and models. Additional
information in this section includes ecological modeling and water
quality data sources.
Section II.A presents a discussion of the effects associated with
oxides of nitrogen and sulfur in the ambient air. The discussion is
organized around the types of effects being considered, including
direct effects of gaseous oxides of nitrogen and sulfur, deposition-
related effects related to acidification and nutrient enrichment, and
other effects such as materials damage, climate-related effects and
mercury methylation.
Section II.B presents a summary and discussion of the risk and
exposure assessment performed for each of the four major effects
categories. The REA uses case studies representing the broad geographic
variability of the impacts from oxides of nitrogen and sulfur to
conclude that there are ongoing adverse effects in many ecosystems from
deposition of oxides of nitrogen and sulfur and that under current
emissions scenarios these effects are likely to continue.
Section II.C presents a discussion of adversity linking ecological
effects to measures that can be used to characterize the extent to
which such effects are reasonably considered to be adverse to public
welfare. This involves consideration of how to characterize adversity
from a public welfare perspective. In so doing, consideration is given
to the concept of ecosystem services, the evidence of effects on
ecosystem services, and how ecosystem services can be linked to
ecological indicators.
Section II.D presents an assessment of the adequacy of the current
oxides of nitrogen and oxides of sulfur secondary standards.
Consideration is given to the adequacy of protection afforded by the
current standards for both direct and deposition-related effects, as
well as to the appropriateness of the fundamental structure and the
basic elements of the current standards for providing protection from
deposition-related effects. Considerations as to the extent to which
deposition-related effects that could reasonably be judged to be
adverse to public welfare are occurring under current conditions which
are allowed by the current standards is also considered. Discussion of
the structures and basic elements of the current NO2 and
SO2 secondary standards and whether they are adequate to
protect against such effects is presented.

[[Page 46091]]

A. Ecological Effects

This section discusses the known or anticipated ecological effects
associated with oxides of nitrogen and sulfur, including the direct
effects of gas-phase exposure to oxides of nitrogen and sulfur (section
II.A.1) and effects associated with deposition-related exposure
(sections II.A.2 and 3). Section II.A. 2 addresses effects related to
acidification of aquatic and terrestrial ecosystems and section II A.3
addresses effects related to nutrient enrichment of aquatic and
terrestrial ecosystems. These sections also address questions about the
nature and magnitude of ecosystem responses to reactive nitrogen and
sulfur deposition, including responses related to acidification,
nutrient depletion, and, in Section II.A 4 the mobilization of toxic
metals in sensitive aquatic and terrestrial ecosystems. The
uncertainties and limitations associated with the evidence of such
effects are also discussed throughout this section.
1. Effects Associated With Gas-Phase Oxides of Nitrogen and Sulfur
Ecological effects on vegetation as discussed in earlier reviews as
well as the ISA can be attributed to gas-phase oxides of nitrogen and
sulfur. Acute and chronic exposures to gaseous pollutants such as
SO2, NO2, nitric oxide (NO), nitric acid
(HNO3) and peroxyacetyl nitrite (PAN) are associated with
negative impacts to vegetation. The current secondary NAAQS were set to
protect against direct damage to vegetation by exposure to gas-phase
oxides of nitrogen and sulfur, such as foliar injury, decreased
photosynthesis, and decreased growth. The following summary is a
concise overview of the known or anticipated effects to vegetation
caused by gas phase nitrogen and sulfur. Most phototoxic effects
associated with gas phase oxides of nitrogen and sulfur occur at levels
well above ambient concentrations observed in the U.S. (US EPA, 2008,
section 3.4.2.4).
a. Nature of Ecosystem Responses to Gas-Phase Nitrogen And Sulfur
The 2008 ISA found that gas phase nitrogen and sulfur are
associated with direct phytotoxic effects (US EPA, 2008, section 4.4).
The evidence is sufficient to infer a causal relationship between
exposure to SO2 and injury to vegetation (US EPA, 2008,
section 4.4.1 and 3.4.2.1). Acute foliar injury to vegetation from
SO2 may occur at levels above the current secondary standard
(3-h average of 0.50 ppm). Effects on growth, reduced photosynthesis
and decreased yield of vegetation are also associated with increased
SO2 exposure concentration and time of exposure.
The evidence is sufficient to infer a causal relationship between
exposure to NO, NO2 and PAN and injury to vegetation (US
EPA, 2008, section 4.4.2 and 3.4.2.2). At sufficient concentrations,
NO, NO2 and PAN can decrease photosynthesis and induce
visible foliar injury to plants. Evidence is also sufficient to infer a
causal relationship between exposure to HNO3 and changes to
vegetation (US EPA, 2008, section 4.4.3 and 3.4.2.3). Phytotoxic
effects of this pollutant include damage to the leaf cuticle in
vascular plants and disappearance of some sensitive lichen species.
b. Magnitude of Ecosystem Response to Gas-Phase Nitrogen And Sulfur
Vegetation in ecosystems near sources of gaseous oxides of nitrogen
and sulfur or where SO2, NO, NO2, PAN and
HNO3 are most concentrated are more likely to be impacted by
these pollutants. Uptake of these pollutants in a plant canopy is a
complex process involving adsorption to surfaces (leaves, stems and
soil) and absorption into leaves (US EPA, 2008, section 3.4.2). The
functional relationship between ambient concentrations of gas phase
oxides of nitrogen and sulfur and specific plant response are impacted
by internal factors such as rate of stomatal conductance and plant
detoxification mechanisms, and external factors including plant water
status, light, temperature, humidity, and pollutant exposure regime (US
EPA, 2008, section 3.4.2).
Entry of gases into a leaf is dependent upon physical and chemical
processes of gas phase as well as to stomatal aperture. The aperture of
the stomata is controlled largely by the prevailing environmental
conditions, such as water availability, humidity, temperature, and
light intensity. When the stomata are closed, resistance to gas uptake
is high and the plant has a very low degree of susceptibility to
injury. Mosses and lichens do not have a protective cuticle barrier to
gaseous pollutants or stomata and are generally more sensitive to
gaseous sulfur and nitrogen than vascular plants (US EPA, 2008, section
3.4.2).
The appearance of foliar injury can vary significantly across
species and growth conditions affecting stomatal conductance in
vascular plants (US EPA, 2009, section 6.4.1). For example, damage to
lichens from SO2 exposure includes decreased photosynthesis
and respiration, damage to the algal component of the lichen, leakage
of electrolytes, inhibition of nitrogen fixation, decreased potassium
(K+) absorption, and structural changes.
The phytotoxic effects of gas phase oxides of nitrogen and sulfur
are dependent on the exposure concentration and duration and species
sensitivity to these pollutants. Effects to vegetation associated with
oxides of nitrogen and sulfur are therefore variable across the U.S.
and tend to be higher near sources of photochemical smog. For example,
SO2 is considered to be the primary factor contributing to
the death of lichens in many urban and industrial areas.
The ISA states there is very limited new research on phytotoxic
effects of NO, NO2, PAN and HNO3 at
concentrations currently observed in the U.S. with the exception of
some lichen species (US EPA, 2008, section 4.4). Past and current
HNO3 concentrations may be contributing to the decline in
lichen species in the Los Angeles basin. Most phytotoxic effects
associated with gas phase oxides of nitrogen and sulfur occur at levels
well above ambient concentrations observed in the U.S. (US EPA, 2008,
section 3.4.2.4).
2. Acidification Effects Associated With Deposition of Oxides of
Nitrogen and Sulfur
Sulfur oxides and nitrogen oxides in the atmosphere undergo a
complex mix of reactions in gaseous, liquid, and solid phases to form
various acidic compounds. These acidic compounds are removed from the
atmosphere through deposition: either wet (e.g., rain, snow), fog or
cloud, or dry (e.g., gases, particles). Deposition of these acidic
compounds to ecosystems can lead to effects on ecosystem structure and
function. Following deposition, these compounds can, in some instances,
unless retained by soil or biota, leach out of the soils in the form of
sulfate (SO4^2-) and nitrate
(NO3^-), leading to the acidification of surface
waters. The effects on ecosystems depend on the magnitude and rate of
deposition, as well as a host of biogeochemical processes occurring in
the soils and water bodies (US EPA, 2009, section 2.1). The chemical
forms of nitrogen that may contribute to acidifying deposition include
both oxidized and reduced chemical species, including reduced forms of
nitrogen (NHx).
When sulfur or nitrogen leaches from soils to surface waters in the
form of SO4^2- or NO3^-, an
equivalent amount of positive cations, or countercharge, is also
transported. This maintains electroneutrality. If the countercharge is
provided by base cations, such as

[[Page 46092]]

calcium (Ca\2+\), magnesium (Mg\2+\), sodium (Na\+\), or K\+\, rather
than hydrogen (H\+\) and dissolved inorganic aluminum, the acidity of
the soil water is neutralized, but the base saturation of the soil
decreases. Continued SO4\2-\ or NO3^-
leaching can deplete the available base cation pool in soil. As the
base cations are removed, continued deposition and leaching of
SO4^2- and/or NO3^- (with
H\+\ and Al\3+\) leads to acidification of soil water, and by
connection, surface water. Introduction of strong acid anions such as
sulfate and nitrate to an already acidic soil, whether naturally or due
to anthropogenic activities, can lead to instantaneous acidification of
waterbodies through direct runoff without any significant change in
base cation saturation. The ability of a watershed to neutralize acidic
deposition is determined by a variety of biogeophysical factors
including weathering rates, bedrock composition, vegetation and
microbial processes, physical and chemical characteristics of soils and
hydrologic flowpaths (US EPA, 2009, section 2.1). Some of these factors
such as vegetation and soil depth are highly variable over small
spatial scales such as meters, but can be aggregated to evaluate
patterns over larger spatial scales. Acidifying deposition of oxides of
nitrogen and sulfur and the chemical and biological responses
associated with these inputs vary temporally. Chronic or long-term
deposition processes in the time scale of years to decades result in
increases in inputs of nitrogen and sulfur to ecosystems and the
associated ecological effects. Episodic or short term (i.e., hours or
days) deposition refers to events in which the level of the acid
neutralizing capacity (ANC) of a lake or stream is temporarily lowered.
In aquatic ecosystems, short-term (i.e., hours or days) episodic
changes in water chemistry can have significant biological effects.
Episodic acidification refers to conditions during precipitation or
snowmelt events when proportionately more drainage water is routed
through upper soil horizons that tend to provide less acid neutralizing
than is passing through deeper soil horizons (US EPA, 2009, section
4.2). In addition, the accumulated sulfate and nitrate in snow packs
can provide a surge of acidic inputs. Some streams and lakes may have
chronic or base flow chemistry that is suitable for aquatic biota, but
may be subject to occasional acidic episodes with deleterious
consequences to sensitive biota.
The following summary is a concise overview of the known or
anticipated effects caused by acidification to ecosystems within the
U.S. Acidification affects both terrestrial and freshwater aquatic
ecosystems.
a. Nature of Acidification-Related Ecosystem Responses
The ISA concluded that deposition of oxides of nitrogen and sulfur
and NHx leads to the varying degrees of acidification of
ecosystems (US EPA, 2008). In the process of acidification,
biogeochemical components of terrestrial and freshwater aquatic
ecosystems are altered in a way that leads to effects on biological
organisms. Deposition to terrestrial ecosystems often moves through the
soil and eventually leaches into adjacent water bodies.
i. Aquatic Ecosystems
The scientific evidence is sufficient to infer a causal
relationship between acidifying deposition and effects on
biogeochemistry and biota in aquatic ecosystems (US EPA, 2008, section
4.2.2). The strongest evidence comes from studies of surface water
chemistry in which acidic deposition is observed to alter sulfate and
nitrate concentrations in surface waters, the sum of base cations, ANC,
dissolved inorganic aluminum and pH (US EPA, 2008, section 3.2.3.2).
The ANC is a key indicator of acidification with relevance to both
terrestrial and aquatic ecosystems. The ANC is useful because it
integrates the overall acid-base status of a lake or stream and
reflects how aquatic ecosystems respond to acidic deposition over time.
There is also a relationship between ANC and the surface water
constituents that directly contribute to or ameliorate acidity-related
stress, in particular, concentrations of hydrogen ion (as pH), Ca\2+\
and aluminum (Al). Moreover, low pH surface waters leach aluminum from
soils, which is quite lethal to fish and other aquatic organisms. In
aquatic systems, there is a direct relationship between ANC and fish
and phyto-zooplankton diversity and abundance.
Low ANC coincides with effects on aquatic systems (e.g., individual
species fitness loss or death, reduced species richness, altered
community structure). At the community level, species richness is
positively correlated with pH and ANC because energy cost in
maintaining physiological homeostasis, growth, and reproduction is high
at low ANC levels. For example, there is a logistic relationship
between fish species richness and ANC class for Adirondack Case Study
Area lakes that indicates the probability of occurrence of an organism
for a given value of ANC. Biota are generally not harmed when ANC
values are >100 microequivalents per liter ([mu]eq/L). The number of
fish species also peaks at ANC values >100 [mu]eq/L. Below 100 [mu]eq/L
ANC, fish fitness and community diversity begin to decline (US EPA,
section 4.2). Specifically at ANC levels between 100 and 50 [mu]eq/L,
the fitness of sensitive species (e.g., brook trout, zooplankton)
begins to decline. When ANC concentrations are <50 [mu]eq/L, they are
generally associated with death or loss of fitness of biota that are
sensitive to acidification.
Consistent and coherent documentation from multiple studies on
various species from all major trophic levels of aquatic systems shows
that geochemical alteration caused by acidification can result in the
loss of acid-sensitive biological species (US EPA, 2008, section
3.2.3.3). This is most often discussed with relation to pH. For
example, in the Adirondacks, of the 53 fish species recorded in
Adirondack lakes about half (26 species) were absent from lakes with pH
below 6.0. Biological effects are linked to changes in water chemistry
including decreases in ANC and pH and increases in inorganic Al
concentration. The direct biological effects are caused by lowered pH
which leads to increased inorganic Al concentrations (US EPA, 2011,
Figures 3-1 and 3-2). While ANC level does not cause direct biological
harm it is a good overall indicator of the risk of acidification (US
EPA, 2011, section 3.1.3).
There are clear associations between ANC, pH and aquatic species
mortality and health which are summarized in section 3.1.1 of the PA.
Significant harm to sensitive aquatic species has been observed at pH
levels below 6. Normal stream pH levels with little to no toxicity
range from 6 to 7 (MacAvoy et al, 1995). Baker et al (1990) observed
that ``lakes with pH less than approximately 6.0 contain significantly
fewer species than lakes with pH levels above 6.0.'' As noted in
Chapter 3, typically at pH <4.5 and an ANC <0 [mu]eq/L, complete to
near-complete loss of many taxa of organisms occur, including fish and
aquatic insect populations, whereas other taxa are reduced to only
acidophilic species. Acid Neutralizing Capacity is a measure of how
much acid can be neutralized in a specific surface water system. An ANC
value of 0 or below means that surface waters have no ability to
neutralize any additional acid inputs.
Additional evidence can help refine the understanding of effects
occurring at pH levels between 4.5 and 6. When pH levels are below 5.6,
relatively lower trout survival rates were observed in the

[[Page 46093]]

Shenandoah National Park. In field observations, when pH levels dropped
to 5, mortality rates went to 100 percent (Bulger et al, 2000). At pH
levels ranging from 5.4 to 5.8, cumulative mortality continues to
increase. Several studies have shown that trout exposed to water with
varying pH levels and fish larvae showed increasing mortality as pH
levels decrease. In one study almost 100 percent mortality was observed
at a pH of 4.5 compared to almost 100 percent survival at a pH of 6.5.
Intermediate pH values (6.0, 5.5) in all cases showed reduced survival
compared with the control (6.5), but not by statistically significant
amounts (US EPA, 2008, section 3.2.3.3).
One important indicator of acid stress is increased fish mortality.
The response of fish to pH is not uniform across species. A number of
synoptic surveys indicated loss of species diversity and absence of
several fish species in the pH range of 5.0 to 5.5. If pH is lower,
there is a greater likelihood that more fish species could be lost
without replacement, resulting in decreased richness and diversity. In
general, populations of salmonids are not found at pH levels less than
5.0, and smallmouth bass (Micropterus dolomieu) populations are usually
not found at pH values less than about 5.2 to 5.5. From Table 3-1, only
one study showed significant mortality effects above a pH of 6, while a
number of studies showed significant mortality when pH levels are at or
below 5.5.
The highest pH level for any of the studies reported in the ISA is
6.0, suggesting that pH above 6.0 is protective against mortality
effects for most species. Most thresholds are in the range of pH of 5.0
to 6.0, which suggests that a target pH should be no lower than 5.0.
Protection against mortality in some recreationally important species
such as lake trout (pH threshold of 5.6) and crappie (pH threshold of
5.5), combined with the evidence of effects on larval and embryo
survival suggests that pH levels greater than 5.5 should be targeted to
provide protection against mortality effects throughout the life stages
of fish.
Non-lethal effects have been observed at pH levels as high as 6. A
study in the Shenandoah National Park found that the condition factor,
a measure of fish health expressed as fish weight/length multiplied by
a scaling constant, is positively correlated with stream pH levels, and
that the condition factor is reduced in streams with a pH of 6.0 (US
EPA, 2008, section 3.2.3.3).
Biodiversity is another indicator of aquatic ecosystem health. A
key study in the Adirondacks found that lakes with a pH of 6.0 had only
half the potential species of fish (27 of 53 potential species). There
is often a positive relationship between pH and number of fish species,
at least for pH values between about 5.0 and 6.5, or ANC values between
about 0 to 100 [mu]eq/L. Such observed relationships are complicated,
however, by the tendency for smaller lakes and streams, having smaller
watersheds, to also support fewer fish species, irrespective of acid-
base chemistry. This pattern may be due to a decrease in the number of
available niches as stream or lake size decreases. Nevertheless, fish
species richness is relatively easily determined and is one of the most
useful indicators of biological effects of surface water acidification.
Changes in stream water pH and ANC also contribute to declines in
taxonomic richness of zooplankton, and macroinvertebrates which are
often sources of food for fish, birds and other animal species in
various ecosystems. These fish may also serve as a source of food and
recreation for humans. Acidification of ecosystems has been shown to
disrupt food web dynamics causing alteration to the diet, breeding
distribution, and reproduction of certain species of birds (US EPA,
2008, section 4.2.2.2. and Table 3-9). For example, breeding
distribution of the common goldeneye (Bucephala clangula), an
insectivorous duck, may be affected by changes in acidifying
deposition. Similarly, decreases in prey diversity and quantity have
been observed to create feeding problems for nesting pairs of loons on
low-pH lakes in the Adirondacks.
ii. Terrestrial Ecosystems
In terrestrial ecosystems, the evidence is sufficient to infer a
causal relationship between acidifying deposition and changes in
biogeochemistry (US EPA, 2008, section 4.2.1.1). The strongest evidence
comes from studies of forested ecosystems, with supportive information
on other plant taxa, including shrubs and lichens (US EPA, 2008,
section 3.2.2.1.). Three useful indicators of chemical changes and
acidification effects on terrestrial ecosystems, showing consistency
and coherence among multiple studies are: soil base saturation, Al
concentrations in soil water, and soil carbon to nitrogen (C:N) ratio
(US EPA, 2008, section 3.2.2.2).
As discussed in the ISA and REA, in soils with base saturation less
than about 15 to 20 percent, exchange chemistry is dominated by Al.
Under these conditions, responses to inputs of sulfuric acid and
HNO3 largely involve the release and mobilization of
dissolved inorganic Al. The effect can be neutralized by weathering
from geologic parent material or base cation exchange. The Ca\2+\ and
Al concentrations in soil water are strongly influenced by soil
acidification and both have been shown to have quantitative links to
tree health, including Al interference with Ca\2+\ uptake and Al
toxicity to roots. Effects of nitrification and associated
acidification and cation leaching have been consistently shown to occur
only in soils with a C:N ratio below about 20 to 25.
Soil acidification caused by acidic deposition has been shown to
cause decreased growth and increased susceptibility to disease and
injury in sensitive tree species. Red spruce (Picea rubens) dieback or
decline has been observed across high elevation areas in the
Adirondack, Green and White mountains. The frequency of freezing injury
to red spruce needles has increased over the past 40 years, a period
that coincided with increased emissions of sulfur and nitrogen oxides
and increased acidifying deposition. Acidifying deposition can
contribute to dieback in sugar maple (Acer saccharum) through depletion
of cations from soil with low levels of available Ca. Grasslands are
likely less sensitive to acidification than forests due to grassland
soils being generally rich in base cations.
iii. Ecosystem Sensitivity
The intersection between current deposition loading, historic
loading and sensitivity defines the ecological vulnerability to the
effects of acidification. Freshwater aquatic and some terrestrial
ecosystems, notably forests, are the ecosystem types which are most
sensitive to acidification. The ISA reports that the principal factor
governing the sensitivity of terrestrial and aquatic ecosystems to
acidification from sulfur and nitrogen deposition is geology
(particularly surficial geology). Geologic formations having low base
cation supply generally underlie the watersheds of acid-sensitive lakes
and streams. Other factors that contribute to the sensitivity of soils
and surface waters to acidifying deposition include topography, soil
chemistry, land use, and hydrologic flowpaths. Episodic and chronic
acidification tends to occur in areas that have base-poor bedrock, high
relief, and shallow soils (US EPA, 2008, section 3.2.4.1).
b. Magnitude of Acidification-Related Ecosystem Responses
Terrestrial and aquatic ecosystems differ in their response to
acidifying

[[Page 46094]]

deposition. Therefore the magnitude of ecosystem response is described
separately for aquatic and terrestrial ecosystems in the following
sections. The magnitude of response refers to both the severity of
effects and the spatial extent of the U.S. which is affected.
i. Aquatic Acidification
Freshwater ecosystem surveys and monitoring in the eastern U.S.
have been conducted by many programs since the mid-1980s, including
EPA's Environmental Monitoring and Assessment Program (EMAP), National
Surface Water Survey (NSWS), Temporally Integrated Monitoring of
Ecosystems (TIME), and Long-term Monitoring (LTM) programs. Based on
analyses of surface water data from these programs, New England, the
Adirondack Mountains, the Appalachian Mountains (northern Appalachian
Plateau and Ridge/Blue Ridge region) and the Upper Midwest contain the
most sensitive lakes and streams (i.e., ANC less than about 50 [mu]eq/
L). Portions of northern Florida also contain many acidic and low-ANC
lakes and streams, although the role of acidifying deposition in this
region is less clear. The western U.S. contains many of the surface
waters most sensitive to potential acidification effects, but with the
exception of the Los Angeles Basin and surrounding areas, the levels of
acidifying deposition are low in most areas. Therefore, acidification
of surface waters by acidic deposition is not as prevalent in the
western U.S., and the extent of chronic surface water acidification
that has occurred in that region to date has likely been very limited
relative to the Eastern U.S. (US EPA, 2008, section 3.2.4.2 and US EPA,
2009, section 4.2.2).
There are a number of species including fish, aquatic insects,
other invertebrates and algae that are sensitive to acidification and
cannot survive, compete or reproduce in acidic waters (US EPA, 2008,
section 3.2.3.3). Decreases in ANC and pH have been shown to contribute
to declines in species richness and declines in abundance of
zooplankton, macroinvertebrates, and fish. Reduced growth rates have
been attributed to acid stress in a number of fish species including
Atlantic salmon (Salmo salar), Chinook salmon (Oncorhynchus
tshawytscha), lake trout (Salvelinus namaycush), rainbow trout
(Oncorhynchis mykiss), brook trout (Salvelinus Fontinalis), and brown
trout (Salmo trutta). In response to small to moderate changes in
acidity, acid-sensitive species are often replaced by other more acid-
tolerant species, resulting in changes in community composition and
richness. The effects of acidification are continuous, with more
species being affected at higher degrees of acidification. At a point,
typically a pH <4.5 and an ANC <0 [mu]eq/L, complete to near-complete
loss of many taxa of organisms occur, including fish and aquatic insect
populations, whereas other taxa are reduced to only acidophilic
species. These changes in taxa composition are associated with the high
energy cost in maintaining physiological homeostasis, growth, and
reproduction at low ANC levels (US EPA, 2008, section 3.2.3.3).
Decreases in species richness related to acidification have been
observed in the Adirondack Mountains and Catskill Mountains of New
York, New England and Pennsylvania, and Virginia. From the sensitive
areas identified by the ISA, further ``case study'' analyses on aquatic
ecosystems in the Adirondack Mountains and Shenandoah National Park
were conducted to better characterize ecological risk associated with
acidification (US EPA, 2009, section 4).
The ANC is the most widely used indicator of acid sensitivity and
has been found in various studies to be the best single indicator of
the biological response and health of aquatic communities in acid-
sensitive systems (Lien et al., 1992; Sullivan et al., 2006; US EPA,
2008). In the REA, surface water trends in SO4^2-
and NO3^- concentrations and ANC levels were
analyzed to affirm the understanding that reductions in deposition
could influence the risk of acidification. The ANC values have been
categorized according to their effects on biota, as shown in the table
below. Monitoring data from TIME/LTM and EMAP programs were assessed
for the years 1990 to 2006, and past, present and future water quality
levels were estimated by both steady-state and dynamic biogeochemical
models.

Table II-1--Ecological Effects Associated With Alternative Levels of
Acid Neutralizing Capacity (ANC)
[Source: USEPA, Acid Rain Program]
------------------------------------------------------------------------

------------------------------------------------------------------------
Category Label ANC Levels and Expected Ecological Effects
------------------------------------------------------------------------
Acute Concern................. <0 [mu]eq/L...... Complete loss of fish
populations is
expected. Planktonic
communities have
extremely low
diversity and are
dominated by
acidophilic taxa.
The numbers of
individuals in
plankton species
that are present are
greatly reduced.
Severe Concern................ 0-20 [mu]eq/L.... Highly sensitive to
episodic
acidification.
During episodes of
high acidifying
deposition, brook
trout populations
may experience
lethal effects. The
diversity and
distribution of
zooplankton
communities decline
sharply.
Elevated Concern.............. 20-50 [mu]eq/L... Fish species richness
is greatly reduced
(i.e., more than
half of expected
species can be
missing). On
average, brook trout
populations
experience sublethal
effects, including
loss of health,
ability to
reproduce, and
fitness. Diversity
and distribution of
zooplankton
communities decline.
Moderate Concern.............. 50-100 [mu]eq/L.. Fish species richness
begins to decline
(i.e., sensitive
species are lost
from lakes). Brook
trout populations
are sensitive and
variable, with
possible sublethal
effects. Diversity
and distribution of
zooplankton
communities also
begin to decline as
species that are
sensitive to
acidifying
deposition are
affected.
Low Concern................... >100 [mu]eq/L.... Fish species richness
may be unaffected.
Reproducing brook
trout populations
are expected where
habitat is suitable.
Zooplankton
communities are
unaffected and
exhibit expected
diversity and
distribution.
------------------------------------------------------------------------

[[Page 46095]]

Studies on fish species richness in the Adirondacks Case Study Area
demonstrated the effect of acidification. Of the 53 fish species
recorded in Adirondack Case Study Area lakes, only 27 species were
found in lakes with a pH <6.0. The 26 species missing from lakes with a
pH <6.0 include important recreational species, such as Atlantic
salmon, tiger trout (Salmo trutta X Salvelinus fontinalis), redbreast
sunfish (Lepomis auritus), bluegill (Lepomis macrochirus), tiger musky
(Esox masquinongy X lucius), walleye (Sander vitreus), alewife (Alosa
pseudoharengus), and kokanee (Oncorhynchus nerka), as well as
ecologically important minnows that are commonly consumed by sport
fish. A survey of 1,469 lakes in the late 1980s found 346 lakes to be
devoid of fish. Among lakes with fish, there was a relationship between
the number of fish species and lake pH, ranging from about one species
per lake for lakes having a pH <4.5 to about six species per lake for
lakes having a pH >6.5. In the Adirondacks, a positive relationship
exists between the pH and ANC in lakes and the number of fish species
present in those lakes (US EPA, 2008, section 3.2.3.4).
Since the mid-1990s, streams in the Shenandoah Case Study Area have
shown slight declines in NO3^- and
SO4^2- concentrations in surface waters. The 2006
concentrations are still above pre-acidification (1860) conditions.
Model of Acidification of Groundwater in Catchments (MAGIC) modeling
predicts surface water concentrations of NO3^- and
SO4^2- are 10- and 32-fold higher, respectively,
in 2006 than in 1860. The estimated average ANC across 60 streams in
the Shenandoah Case Study Area is 57.9 [mu]eq/L ( 4.5
[mu]eq/L). Fifty-five percent of all monitored streams in the
Shenandoah Case Study Area have a current risk of Elevated, Severe, or
Acute. Of the 55 percent, 18 percent are chronically acidic today (US
EPA, 2009, section 4.2.4.3).
Based on a deposition scenario for this study area that maintains
current emission levels from 2020 to 2050, the simulation forecast
indicates that a large number of streams would still have Elevated to
Acute problems with acidity in 2050.
Biological effects of increased acidification documented in the
Shenandoah Case Study Area include a decrease in the condition factor
in blacknose dace and a decrease in fish biodiversity associated with
decreasing stream ANC. On average, the fish species richness is lower
by one fish species for every 21 [mu]eq/L decrease in ANC in Shenandoah
National Park streams (US EPA, 2008, section 3.2.3.4).
ii. Terrestrial Acidification
The ISA identified a variety of indicators that can be used to
measure the effects of acidification in soils. Most effects of
terrestrial acidification are observed in sensitive forest ecosystem in
the U.S. Tree health has been linked to the availability of base
cations (BC) in soil (such as Ca\2+\, Mg\2+\ and K\+\), as well as soil
aluminum (Al) content. Tree species show a range of sensitivities to
Ca/Al and BC/Al soil molar ratios, therefore these are good chemical
indicators because they directly relate to the biological effects.
Critical BC/Al molar ratios for a large variety of tree species ranged
from 0.2 to 0.8. This range is similar to critical ratios of Ca/Al.
Plant toxicity or nutrient antagonism was reported to occur at Ca/Al
molar ratios ranging from 0.2 to 2.5 (US EPA, 2009).
There has been no systematic national survey of terrestrial
ecosystems to determine the extent and distribution of terrestrial
ecosystem sensitivity to the effects of acidifying deposition. However,
one preliminary national evaluation estimated that ~15 percent of
forest ecosystems in the U.S. exceed the estimated critical load based
on soil ANC leaching for sulfur and nitrogen deposition by >250 eq/ha/
yr (McNulty et al., 2007). Forests of the Adirondack Mountains of New
York, Green Mountains of Vermont, White Mountains of New Hampshire, the
Allegheny Plateau of Pennsylvania and high-elevation forest ecosystems
in the southern Appalachians are the regions most sensitive to
terrestrial acidification effects from acidifying deposition (US EPA,
2008, section 3.2.4.2). While studies show some recovery of surface
waters, there are widespread measurements of ongoing depletion of
exchangeable base cations in forest soils in the northeastern U.S.
despite recent decreases in acidifying deposition, indicating a slow
recovery time.
In the REA, a critical load analysis was performed for sugar maple
and red spruce forests in the eastern U.S. by using BC/Al ratio in
acidified forest soils as an indicator to assess the impact of nitrogen
and sulfur deposition on tree health. These are the two most commonly
studied tree species in North America for effects of acidification. At
a BC/Al ratio of 1.2, red spruce growth can be decreased by 20 percent.
Sugar maple growth can be decreased by 20 percent at a BC/Al ratio of
0.6 (US EPA, 2009, section 4.4). The REA analysis determined the health
of at least a portion of the sugar maple and red spruce growing in the
U.S. may have been compromised with acidifying total nitrogen and
sulfur deposition. Specifically, total nitrogen and sulfur deposition
levels exceeded three selected critical loads for tree growth in 3
percent to 75 percent of all sugar maple plots across 24 states--that
is, it exceeded the highest (least stringent) of the three critical
loads in 3 percent of plots, and the lowest (most stringent) in 75
percent of plots. For red spruce, total nitrogen and sulfur deposition
levels exceeded three selected critical loads in 3 percent to 36
percent of all red spruce plots across eight states (US EPA, 2009,
section 4.4).
c. Key Uncertainties Associated With Acidification
There are different levels of uncertainty associated with
relationships between deposition, ecological effects and ecological
indicators. In Chapter 7 of the REA, the case study analyses associated
with each targeted effect area were synthesized by identifying the
strengths, limitations, and uncertainties associated with the available
data, modeling approach, and relationship between the selected
ecological indicator and atmospheric deposition as described by the
ecological effect function (US EPA, 2009, Figure 1-1). A further
discussion of uncertainty in aquatic and terrestrial ecosystems is
presented below. The key uncertainties were characterized as follows to
evaluate the strength of the scientific basis for setting a national
standard to protect against a given effect (US EPA, 2009, section 7):
(1) Data Availability: High, medium or low quality. This criterion
is based on the availability and robustness of data sets, monitoring
networks, availability of data that allows for extrapolation to larger
assessment areas and input parameters for modeling and developing the
ecological effect function. The scientific basis for the ecological
indicator selected is also incorporated into this criterion.
(2) Modeling Approach: High, fairly high, intermediate, or low
confidence. This value is based on the strengths and limitations of the
models used in the analysis and how accepted they are by the scientific
community for their application in this analysis.
(3) Ecological Effect Function: High, fairly high, intermediate or
low confidence. This ranking is based on how well the ecological effect
function describes the relationship between atmospheric deposition and
the ecological indicator of an effect.

[[Page 46096]]

i. Aquatic Acidification
The REA concludes that the available data are robust and considered
high quality. There is high confidence about the use of these data and
their value for extrapolating to a larger regional population of lakes.
The EPA TIME/LTM network represents a source of long-term,
representative sampling. Data on sulfate concentrations, nitrate
concentrations and ANC from 1990 to 2006 used for this analysis as well
as EPA EMAP and Regional Environmental Monitoring and Assessment
Program (REMAP) surveys, provide considerable data on surface water
trends.
There is fairly high confidence associated with modeling and input
parameters. Uncertainties in water quality estimates (i.e., ANC) from
MAGIC were derived from multiple site calibrations. Pre-acidification
refers to retrospective modeling to estimate water quality conditions
before man-made contributions of acidifying inputs. The models are
evaluated under current conditions to determine how well they replicate
observed ANC values. The 95 percent confidence interval for pre-
acidification of lakes was an average of 15 [micro]eq/L difference in
ANC concentrations, or 10 percent, and 8 [micro]eq/L, or 5 percent, for
streams (US EPA, 2009, section 7.1.2). The use of the critical load
model to estimate aquatic critical loads is limited by the
uncertainties associated with runoff and surface water measurements and
in estimating the catchment supply of base cations from the weathering
of bedrock and soils (McNulty et al., 2007).
ii. Terrestrial Acidification
The available data used to quantify the targeted effect of
terrestrial acidification are robust and considered high quality. The
U.S. Forest Service-Kane Experimental Forest and significant amounts of
research work in the Allegheny Plateau have produced extensive, peer-
reviewed data sets. Sugar maple and red spruce were the focus of the
REA since they are demonstrated to be negatively affected by soil
available Ca\2+\ depletion and high concentrations of available Al, and
occur in areas that receive high acidifying deposition. There is high
confidence about the use of the REA terrestrial acidification data and
their value for extrapolating to a larger regional population of
forests.
There is high confidence associated with the models, input
parameters, and assessment of uncertainty used in the case study for
terrestrial acidification. The Simple Mass Balance (SMB) model, a
commonly used and widely applied approach for estimating critical
loads, was used in the REA analysis (US EPA, 2008, section 7.2.2).
There is fairly high confidence associated with the ecological effect
function developed for terrestrial acidification (US EPA, 2009, section
7.2.3).
3. Nutrient Enrichment Effects Associated With Deposition of Oxides of
Nitrogen
The following summary is a concise overview of the known or
anticipated effects caused by nitrogen nutrient enrichment to
ecosystems within the United States. Nutrient-enrichment affects
terrestrial, freshwater and estuarine ecosystems. Nitrogen deposition
is a major source of anthropogenic nitrogen. For many terrestrial and
freshwater ecosystems other sources of nitrogen including fertilizer
and waste treatment are greater than deposition. Nitrogen deposition
often contributes to nitrogen-enrichment effects in estuaries, but does
not drive the effects since other sources of nitrogen greatly exceed
nitrogen deposition. Both oxides of nitrogen and NHX
contribute to nitrogen deposition. For the most part, nitrogen effects
on ecosystems do not depend on whether the nitrogen is in oxidized or
reduced form. Thus, this summary focuses on the effects of nitrogen
deposition in total.
a. Nature of Nutrient Enrichment-Related Ecosystem Responses
The ISA found that deposition of nitrogen, including oxides of
nitrogen and NHX, leads to the nitrogen enrichment of
ecosystems (US EPA 2008). In the process of nitrogen enrichment,
biogeochemical components of terrestrial and freshwater aquatic
ecosystems are altered in a way that leads to effects on biological
organisms.
i. Aquatic Ecosystems
In freshwater ecosystems, the evidence is sufficient to infer a
causal relationship between nitrogen deposition and the alteration of
biogeochemical cycling in freshwater aquatic ecosystems (US EPA, 2008,
section 3.3.2.3). Nitrogen deposition is the main source of nitrogen
enrichment to headwater streams, lower order streams and high elevation
lakes. The most common chemical indicators that were studied included
NO3^2- and dissolved inorganic nitrogen (DIN)
concentration in surface waters as well as the ratio of chlorophyll a
to total phosphorus. Elevated surface water NO3^-
concentrations occur in both the eastern and western U.S. Studies
report a significant correlation between nitrogen deposition and lake
biogeochemistry by identifying a correlation between wet deposition and
DIN and the ratio of chlorophyll a to total phosphate. Recent evidence
provides examples of lakes and streams that are limited by nitrogen and
show signs of eutrophication in response to nitrogen addition.
The evidence is sufficient to infer a causal relationship between
nitrogen deposition and the alteration of species richness, species
composition and biodiversity in freshwater aquatic ecosystems (US EPA,
2008, section 3.3.5.3). Increased nitrogen deposition can cause a shift
in community composition and reduce algal biodiversity, especially in
sensitive oligotrophic lakes.
In the ISA, the evidence is sufficient to infer a causal
relationship between nitrogen deposition and the biogeochemical cycling
of nitrogen and carbon in estuaries (US EPA, 2008, section 4.3.4.1 and
3.3.2.3). In general, estuaries tend to be nitrogen-limited, and many
currently receive high levels of nitrogen input from human activities
(US EPA, 2009, section 5.1.1). It is unknown if atmospheric deposition
alone is sufficient to cause eutrophication; however, the contribution
of atmospheric nitrogen deposition to total nitrogen load is calculated
for some estuaries and can be >40 percent (US EPA, 2009, section
5.1.1).
The evidence is sufficient to infer a causal relationship between
nitrogen deposition and the alteration of species richness, species
composition and biodiversity in estuarine ecosystems (US EPA, 2008,
section 4.3.4.2 and 3.3.5.4). Atmospheric and non-atmospheric sources
of nitrogen contribute to increased phytoplankton and algal
productivity, leading to eutrophication. Shifts in community
composition, reduced hypolimnetic dissolved oxygen (DO), decreases in
biodiversity, and mortality of submerged aquatic vegetation are
associated with increased N deposition in estuarine systems.
ii. Terrestrial Ecosystems
The evidence is sufficient to infer a causal relationship between
nitrogen deposition and the alteration of biogeochemical cycling in
terrestrial ecosystems (US EPA, 2008, section 4.3.1.1 and 3.3.2.1).
This is supported by numerous observational, deposition gradient and
field addition experiments in sensitive ecosystems. The leaching of
NO3^- in soil drainage waters and the export of
NO3^- in stream water were identified as two of
the primary

[[Page 46097]]

indictors of nitrogen enrichment. Several nitrogen-addition studies
indicate that NO3^- leaching is induced by chronic
additions of nitrogen. Studies identified in the ISA found that surface
water NO3^- concentrations exceeded 1 [mu]eq/L in
watersheds receiving about 9 to 13 kg N/ha/yr of atmospheric nitrogen
deposition. Nitrogen deposition disrupts the nutrient balance of
ecosystems with numerous biogeochemical effects. The chemical
indicators that are typically measured include
NO3^- leaching, soil C:N ratio, rates of nitrogen
mineralization, nitrification, denitrification, foliar nitrogen
concentration, and soil water NO3^- and
NH4⁺ concentrations. Note that nitrogen
saturation (nitrogen leaching from ecosystems) does not need to occur
to cause effects. Substantial leaching of NO3^-
from forest soils to stream water can acidify downstream waters,
leading to effects described in the previous section on aquatic
acidification. Due to the complexity of interactions between the
nitrogen and carbon cycling, the effects of nitrogen on carbon budgets
(quantified input and output of carbon to the ecosystem) are variable.
Regional trends in net ecosystem productivity (NEP) of forests (not
managed for silviculture) have been estimated through models based on
gradient studies and meta-analysis. Atmospheric nitrogen deposition has
been shown to cause increased litter accumulation and carbon storage in
above-ground woody biomass. In the West, this has lead to increased
susceptibility to more severe fires. Less is known regarding the
effects of nitrogen deposition on carbon budgets of non-forest
ecosystems.
The evidence is sufficient to infer a causal relationship between
nitrogen deposition on the alteration of species richness, species
composition and biodiversity in terrestrial ecosystems (US EPA, 2008,
section 4.3.1.2). Some organisms and ecosystems are more sensitive to
nitrogen deposition and effects of nitrogen deposition are not observed
in all habitats. The most sensitive terrestrial taxa to nitrogen
deposition are lichens. Empirical evidence indicates that lichens in
the U.S. are affected by deposition levels as low as 3 kg N/ha/yr.
Alpine ecosystems are also sensitive to nitrogen deposition; changes in
an individual species (Carex rupestris) were estimated to occur at
deposition levels near 4 kg N/ha/yr and modeling indicates that
deposition levels near 10 kg N/ha/yr alter plant community assemblages.
In several grassland ecosystems, reduced species diversity and an
increase in non-native, invasive species are associated with nitrogen
deposition.
iii. Ecosystem Sensitivity to Nutrient Enrichment
The numerous ecosystem types that occur across the U.S. have a
broad range of sensitivity to nitrogen deposition (US EPA, 2008, Table
4-4). Increased deposition to nitrogen-limited ecosystems can lead to
production increases that may be either beneficial or adverse depending
on the system and management goals.
Organisms in their natural environment are commonly adapted to a
specific regime of nutrient availability. Change in the availability of
one important nutrient, such as nitrogen, may result in an imbalance in
ecological stoichiometry, with effects on ecosystem processes,
structure and function. In general, nitrogen deposition to terrestrial
ecosystems causes accelerated growth rates in some species deemed
desirable in commercial forests but may lead to altered competitive
interactions among species and nutrient imbalances, ultimately
affecting biodiversity. The onset of these effects occurs with nitrogen
deposition levels as low as 3 kg N/ha/yr in sensitive terrestrial
ecosystems to nitrogen deposition. In aquatic ecosystems, nitrogen that
is both leached from the soil and directly deposited to the water
surface can pollute the surface water. This causes alteration of the
diatom community at levels as low as 1.5 kg N/ha/yr in sensitive
freshwater ecosystems.
The degree of ecosystem effects lies at the intersection of
nitrogen loading and nitrogen-sensitivity. Nitrogen-sensitivity is
predominately driven by the degree to which growth is limited by
nitrogen availability. Grasslands in the western U.S. are typically
nitrogen-limited ecosystems dominated by a diverse mix of perennial
forbs and grass species. A meta-analysis discussed in the ISA (US EPA,
2008, section 3.3.3), indicated that nitrogen fertilization increased
aboveground growth in all non-forest ecosystems except for deserts. In
other words, almost all terrestrial ecosystems are nitrogen-limited and
will be altered by the addition of anthropogenic nitrogen. Likewise, a
freshwater lake or stream must be nitrogen-limited to be sensitive to
nitrogen-mediated eutrophication. There are many examples of fresh
waters that are nitrogen-limited or nitrogen and phosphorous (P) co-
limited (US EPA, 2008, section 3.3.3.2). A large dataset meta-analysis
discussed in the ISA (US EPA, 2008, section 3.3.3.2), found that
nitrogen-limitation occurred as frequently as phosphorous-limitation in
freshwater ecosystems. Additional factors that govern the sensitivity
of ecosystems to nutrient enrichment from nitrogen deposition include
rates and form of nitrogen deposition, elevation, climate, species
composition, plant growth rate, length of growing season, and soil
nitrogen retention capacity (US EPA, 2008, section 4.3). Less is known
about the extent and distribution of the terrestrial ecosystems in the
U.S. that are most sensitive to the effects of nutrient enrichment from
atmospheric nirogen deposition compared to acidification.
Because the productivity of estuarine and near shore marine
ecosystems is generally limited by the availability of nitrogen, they
are susceptible to the eutrophication effect of nitrogen deposition (US
EPA, 2008, section 4.3.4.1). A recent national assessment of eutrophic
conditions in estuaries found the most eutrophic estuaries were
generally those that had large watershed-to-estuarine surface area,
high human population density, high rainfall and runoff, low dilution
and low flushing rates. In the REA, the National Oceanic and
Atmospheric Administration's (NOAA) National Estuarine Eutrophication
Assessment (NEEA) assessment tool, Assessment of Estuarine Tropic
Status (ASSETS) categorical Eutrophication Index (EI) was used to
evaluate eutrophication due to atmospheric loading of nitrogen. The
ASSETS EI is an estimation of the likelihood that an estuary is
experiencing eutrophication or will experience eutrophication based on
five ecological indicators: Chlorophyll a, macroalgae, dissolved
oxygen, nuisance/toxic algal blooms and submerged aquatic vegetation
(SAV).
In the REA, two regions were selected for case study analysis using
ASSETS EI, the Chesapeake Bay and Pamlico Sound. Both regions received
an ASSETS EI rating of Bad indicating that the estuary had moderate to
high pressure due to overall human influence and a moderate high to
high eutrophic condition (US EPA, 2009, sections 5.2.4.1 and 5.2.4.2).
These results were then considered with SPAtially Referenced Regression
on Watershed Attributes (SPARROW) modeling to develop a response curve
to examine the role of atmospheric nitrogen deposition in achieving a
desired decrease in load. To change the Neuse River Estuary's EI score
from Bad to Poor not only must 100 percent of the total atmospheric
nitrogen deposition be eliminated, but considerably more nitrogen from
other sources as well must be controlled (US EPA, 2009, section
5.2.7.2). In the Potomac River estuary, a 78 percent

[[Page 46098]]

decrease of total nitrogen could move the EI score from Bad to Poor (US
EPA, 2009, section 5.2.7.1). The results of this analysis indicated
decreases in atmospheric deposition alone could not eliminate coastal
eutrophication problems due to multiple non-atmospheric nitrogen inputs
(US EPA, 2009, section 7.3.3). However, the somewhat arbitrary
discreteness of the EI scale can mask the benefits of decreases in
nitrogen between categories.
In general, estuaries tend to be nitrogen-limited, and many
currently receive high levels of nitrogen input from human activities
to cause eutrophication. As reported in the ISA (US EPA, 2008, section
3.2.2.2), atmospheric nitrogen loads to estuaries in the U.S. are
estimated to range from 2 to 8 percent for Guadalupe Bay, Texas on the
lowest end to as high as 72 percent for St. Catherines-Sapelo estuary,
Georgia. The Chesapeake Bay is an example of a large, well-studied and
severely eutrophic estuary that is calculated to receive as much as 30
percent of its total nitrogen load from the atmosphere.
b. Magnitude of Ecosystem Responses
i. Aquatic Ecosystems
The magnitude of ecosystem response may be thought of on two time
scales, current conditions and how ecosystems have been altered since
the onset of anthropogenic nitrogen deposition. As noted previously,
studies found that nitrogen-limitation occurs as frequently as
phosphorous-limitation in freshwater ecosystems (US EPA, 2008, section
3.3.3.2). Recently, a comprehensive study of available data from the
northern hemisphere surveys of lakes along gradients of nitrogen
deposition show increased inorganic nitrogen concentration and
productivity to be correlated with atmospheric nitrogen deposition. The
results are unequivocal evidence of nitrogen limitation in lakes with
low ambient inputs of nitrogen, and increased nitrogen concentrations
in lakes receiving nitrogen solely from atmospheric nitrogen
deposition. It has been suggested that most lakes in the northern
hemisphere may have originally been nitrogen-limited, and that
atmospheric nitrogen deposition has changed the balance of nitrogen and
phosphorous in lakes.
Available data suggest that the increases in total nitrogen
deposition do not have to be large to elicit an ecological effect. For
example, a hindcasting exercise determined that the change in Rocky
Mountain National Park lake algae that occurred between 1850 and 1964
was associated with an increase in wet nitrogen deposition that was
only about 1.5 kg N/ha. Similar changes inferred from lake sediment
cores of the Beartooth Mountains of Wyoming also occurred at about 1.5
kg N/ha deposition. Pre-industrial inorganic nitrogen deposition is
estimated to have been only 0.1 to 0.7 kg N/ha based on measurements
from remote parts of the world. In the western U.S., pre-industrial, or
background, inorganic nitrogen deposition was estimated by to range
from 0.4 to 0.7 kg N/ha/yr.
Eutrophication effects from nitrogen deposition are most likely to
be manifested in undisturbed, low nutrient surface waters such as those
found in the higher elevation areas of the western U.S. The most severe
eutrophication from nitrogen deposition effects is expected downwind of
major urban and agricultural centers. High concentrations of lake or
streamwater NO3^-, indicative of ecosystem
saturation, have been found at a variety of locations throughout the
U.S., including the San Bernardino and San Gabriel Mountains within the
Los Angeles Air Basin, the Front Range of Colorado, the Allegheny
mountains of West Virginia, the Catskill Mountains of New York, the
Adirondack Mountains of New York, and the Great Smoky Mountains in
Tennessee (US EPA, 2008, section 3.3.8).
In contrast to terrestrial and freshwater systems, atmospheric
nitrogen load to estuaries contributes to the total load but does not
necessarily drive the effects since other combined sources of nitrogen
often greatly exceed nitrogen deposition. In estuaries, nitrogen-
loading from multiple anthropogenic and non-anthropogenic pathways
leads to water quality deterioration, resulting in numerous effects
including hypoxic zones, species mortality, changes in community
composition and harmful algal blooms that are indicative of
eutrophication. The following summary is a concise overview of the
known or anticipated effects of nitrogen enrichment on estuaries within
the U.S.
There is a scientific consensus (US EPA, 2008, section 4.3.4) that
nitrogen-driven eutrophication in shallow estuaries has increased over
the past several decades and that the environmental degradation of
coastal ecosystems due to nitrogen, phosphorus, and other inputs is now
a widespread occurrence. For example, the frequency of phytoplankton
blooms and the extent and severity of hypoxia have increased in the
Chesapeake Bay and Pamlico estuaries in North Carolina and along the
continental shelf adjacent to the Mississippi and Atchafalaya rivers'
discharges to the Gulf of Mexico.
A recent national assessment of eutrophic conditions in estuaries
found that 65 percent of the assessed systems had moderate to high
overall eutrophic conditions. Most eutrophic estuaries occurred in the
mid-Atlantic region and the estuaries with the lowest degree of
eutrophication were in the North Atlantic. Other regions had mixtures
of low, moderate, and high degrees of eutrophication (US EPA, 2008,
section 4.3.4.3).
The mid-Atlantic region is the most heavily impacted area in terms
of moderate or high loss of submerged aquatic vegetation due to
eutrophication (US EPA, 2008, section 4.3.4.2). Submerged aquatic
vegetation is important to the quality of estuarine ecosystem habitats
because it provides habitat for a variety of aquatic organisms, absorbs
excess nutrients, and traps sediments (US EPA, 2008, section 4.3.4.2).
It is partly because many estuaries and near-coastal marine waters are
degraded by nutrient enrichment that they are highly sensitive to
potential negative impacts from nitrogen addition from atmospheric
deposition.
ii. Terrestrial Ecosystems
Little is known about the full extent and distribution of the
terrestrial ecosystems in the U.S. that are most sensitive to impacts
caused by nutrient enrichment from atmospheric nitrogen deposition. As
previously stated, most terrestrial ecosystems are nitrogen-limited,
therefore they are sensitive to perturbation caused by nitrogen
additions (US EPA, 2008, section 4.3.1). Effects are most likely to
occur where areas of relatively high atmospheric N deposition intersect
with nitrogen-limited plant communities. The alpine ecosystems of the
Colorado Front Range, chaparral watersheds of the Sierra Nevada, lichen
and vascular plant communities in the San Bernardino Mountains and the
Pacific Northwest, and the southern California coastal sage scrub (CSS)
community are among the most sensitive terrestrial ecosystems. There is
growing evidence (US EPA, 2008, section 4.3.1.2) that existing
grassland ecosystems in the western U.S. are being altered by elevated
levels of N inputs, including inputs from atmospheric deposition.
In the eastern U.S., the degree of nitrogen saturation of the
terrestrial ecosystem is often assessed in terms of the degree of
NO3^- leaching from watershed soils into ground
water or surface water. Studies have estimated the number of surface
waters at different

[[Page 46099]]

stages of saturation across several regions in the eastern U.S. Of the
85 northeastern watersheds examined 60 percent were in Stage 1 or Stage
2 of nitrogen saturation on a scale of 0 (background or pretreatment)
to 3 (visible decline). Of the northeastern sites for which adequate
data were available for assessment, those in Stage 1 or 2 were most
prevalent in the Adirondack and Catskill Mountains. Effects on
individual plant species have not been well studied in the U.S. More is
known about the sensitivity of particular plant communities. Based
largely on results obtained in more extensive studies conducted in
Europe, it is expected that the more sensitive terrestrial ecosystems
include hardwood forests, alpine meadows, arid and semi-arid lands, and
grassland ecosystems (US EPA, 2008, section 3.3.5).
The REA used published research results (US EPA, 2009, section
5.3.1 and US EPA, 2008, Table 4.4) to identify meaningful ecological
benchmarks associated with different levels of atmospheric nitrogen
deposition. These are illustrated in Figure 3-4 of the PA. The
sensitive areas and ecological indicators identified by the ISA were
analyzed further in the REA to create a national map that illustrates
effects observed from ambient and experimental atmospheric nitrogen
deposition loads in relation to Community Multi-scale Air Quality
(CMAQ) 2002 modeling results and National Atmospheric Deposition
Program (NADP) monitoring data. This map, reproduced in Figure 3-5 of
the PA, depicts the sites where empirical effects of terrestrial
nutrient enrichment have been observed and site proximity to elevated
atmospheric nitrogen deposition.
Based on information in the ISA and initial analysis in the REA,
further case study analyses on terrestrial nutrient enrichment of
ecosystems were developed for the CS community and Mixed Conifer Forest
(MCF) (US EPA, 2009). Geographic information systems (GIS) analysis
supported a qualitative review of past field research to identify
ecological benchmarks associated with CSS and mycorrhizal communities,
as well as MCF nutrient-sensitive acidophyte lichen communities, fine-
root biomass in Ponderosa pine, and leached nitrate in receiving
waters.
The ecological benchmarks that were identified for the CSS and the
MCF communities are included in the suite of benchmarks identified in
the ISA (US EPA, 2008, section 3.3). There are sufficient data to
confidently relate the ecological effect to a loading of atmospheric
nitrogen. For the CSS community, the following ecological benchmarks
were identified:

(1) 3.3 kg N/ha/yr--the amount of nitrogen uptake by a vigorous stand
of CSS; above this level, nitrogen may no longer be limiting
(2) 10 kg N/ha/yr--mycorrhizal community changes

For the MCF community, the following ecological benchmarks were
identified:

(1) 3.1 kg N/ha/yr--shift from sensitive to tolerant lichen species
(2) 5.2 kg N/ha/yr--dominance of the tolerant lichen species
(3) 10.2 kg N/ha/yr--loss of sensitive lichen species
(4) 17 kg N/ha/yr--leaching of nitrate into streams.

These benchmarks, ranging from 3.1 to 17 kg N/ha/yr, were compared
to 2002 CMAQ/NADP data to discern any associations between atmospheric
deposition and changing communities. Evidence supports the finding that
nitrogen alters CSS and MCF communities. Key findings include the
following: 2002 CMAQ/NADP nitrogen deposition data show that the 3.3 kg
N/ha/yr benchmark has been exceeded in more than 93 percent of CSS
areas (654,048 ha). These deposition levels are a driving force in the
degradation of CSS communities. Although CSS decline has been observed
in the absence of fire, the contributions of deposition and fire to the
CSS decline require further research. The CSS is fragmented into many
small parcels, and the 2002 CMAQ/NADP 12-km grid data are not fine
enough to fully validate the relationship between CSS distribution,
nitrogen deposition, and fire. The 2002 CMAQ/NADP nitrogen deposition
data exceeds the 3.1 kg N/ha/yr benchmark in more than 38 percent
(1,099,133 ha) of MCF areas, and nitrate leaching has been observed in
surface waters. Ozone effects confound nitrogen effects on MCF
acidophyte lichen, and the interrelationship between fire and nitrogen
cycling requires additional research.
c. Key Uncertainties Associated With Nutrient Enrichment
There are different levels of uncertainty associated with
relationships between deposition, ecological effects and ecological
indicators. The criteria used in the REA to evaluate the degree of
confidence in the data, modeling and ecological effect function are
detailed in chapter 7 of the REA. Below is a discussion of uncertainty
relating aquatic and terrestrial ecosystems to nutrient enrichment
effects.
i. Aquatic Ecosystems
The approach for assessing atmospheric contributions to total
nitrogen loading in the REA was to consider the main-stem river to an
estuary (including the estuary) rather than an entire estuary system or
bay. The biological indicators used in the NOAA ASSETS EI required the
evaluation of many national databases including the US Geological
Survey National Water Quality Assessment (NAWQA) files, EPA's STORage
and RETrieval (STORET) database, NOAA's Estuarine Drainage Areas data
and EPA's water quality standards nutrient criteria for rivers and
lakes (US EPA, 2009, Appendix 6 and Table 1.2.-1). Both the SPARROW
modeling for nitrogen loads and assessment of estuary conditions under
NOAA ASSETS EI, have been applied on a national scale. The REA
concludes that the available data are medium quality with intermediate
confidence about the use of these data and their values for
extrapolating to a larger regional area (US EPA, 2009, section 7.3.1).
Intermediate confidence is associated with the modeling approach using
ASSETS EI and SPARROW. The REA states there is low confidence with the
ecological effect function due to the results of the analysis which
indicated that reductions in atmospheric deposition alone could not
solve coastal eutrophication problems due to multiple non-atmospheric
nitrogen inputs (US EPA, 2009, section 7.3.3).
ii. Terrestrial Ecosystems
Ecological thresholds are identified for CSS and MCF areas and
these data are considered to be of high quality, however, the ability
to extrapolate these data to larger regional areas is limited (US EPA,
2009, section 7.4.1). No quantitative modeling was conducted or
ecological effect function developed for terrestrial nutrient
enrichment reflecting the uncertainties associated with these
depositional effects.
4. Other Ecological Effects
It is stated in the ISA (US EPA, 2008, section 3.4.1 and 4.5) that
mercury is a highly neurotoxic contaminant that enters the food web as
a methylated compound, methylmercury (MeHg). Mercury is principally
methylated by sulfur-reducing bacteria and can be taken up by
microorganisms, zooplankton and macroinvertebrates. The contaminant is
concentrated in higher trophic levels, including fish eaten by humans.
Experimental evidence has established that only inconsequential amounts
of MeHg can

[[Page 46100]]

be produced in the absence of sulfate. Once MdHg is present, other
variables influence how much accumulates in fish, but elevated mercury
levels in fish can only occur where substantial amounts of MeHg are
present. Current evidence indicates that in watersheds where mercury is
present, increased oxides of sulfur deposition very likely results in
additional production of MeHg which leads to greater accumulation of
MeHg concentrations in fish. With respect to sulfur deposition and
mercury methylation, the final ISA determined that ``[t]he evidence is
sufficient to infer a causal relationship between sulfur deposition and
increased mercury methylation in wetlands and aquatic environments.''
The production of meaningful amounts of MeHg requires the presence
of SO4^2- and mercury, and where mercury is
present, increased availability of SO4^2- results
in increased production of MeHg. There is increasing evidence on the
relationship between sulfur deposition and increased methylation of
mercury in aquatic environments; this effect occurs only where other
factors are present at levels within a range to allow methylation. The
production of MeHg requires the presence of SO4^2-
and mercury, but the amount of MeHg produced varies with oxygen
content, temperature, pH, and supply of labile organic carbon (US EPA,
2008, section 3.4). In watersheds where changes in sulfate deposition
did not produce an effect, one or several of those interacting factors
were not in the range required for meaningful methylation to occur (US
EPA, 2008, section 3.4). Watersheds with conditions known to be
conducive to mercury methylation can be found in the northeastern U.S.
and southeastern Canada.
While the relationship between sulfur and MeHg production was
concluded to be causal in the ISA, the REA concluded that there was
insufficient evidence to quantify the relationship between sulfur and
MeHg. Therefore only a qualitative assessment was included in chapter 6
of the REA. The PA was then unable to make a determination as to the
adequacy of the existing SO2 standards in protecting against
welfare effects associated with increased mercury methylation.

B. Risk and Exposure Assessment

The risk and exposure assessment conducted for the current review
was developed to describe potential risk from current and future
deposition of oxides of nitrogen and sulfur to sensitive ecosystems.
The case study analyses in the REA show that there is confidence that
known or anticipated adverse ecological effects are occurring under
current ambient loadings of nitrogen and sulfur in sensitive ecosystems
across the U.S. An overview of the material covered in the REA, a
summary of the key findings from the air quality analyses,
acidification and nutrient enrichment case studies, and general
conclusions from evaluating additional welfare effects, are presented
below.
1. Overview of the Risk and Exposure Assessment
The REA evaluates the relationships between atmospheric
concentrations, deposition, biologically relevant exposures, targeted
ecosystem effects, and ecosystem services. To evaluate the nature and
magnitude of adverse effects associated with deposition, the REA also
examines various ways to quantify the relationships between air quality
indicators, deposition of biologically available forms of nitrogen and
sulfur, ecologically relevant indicators relating to deposition,
exposure and effects on sensitive receptors, and related effects
resulting in changes in ecosystem structure and services. The intent is
to determine the exposure metrics that incorporate the temporal
considerations (i.e., biologically relevant timescales), pathways, and
ecologically relevant indicators necessary to determine the effects on
these ecosystems. To the extent feasible, the REA evaluates the overall
load to the system for nitrogen and sulfur, as well as the variability
in ecosystem responses to these pollutants. It also evaluates the
contributions of atmospherically deposited nitrogen and sulfur
individually relative to the combined atmospheric loadings of both
elements together.. Since oxidized nitrogen is the listed criteria
pollutant (currently measured by the ambient air quality indicator
NO2) for the atmospheric contribution to total nitrogen, the
REA examines the contribution of nitrogen oxides to total reactive
nitrogen in the atmosphere, relative to the contributions of reduced
forms of nitrogen (e.g., ammonia, ammonium), to ultimately assess how a
meaningful secondary NAAQS might be structured.
The REA focuses on ecosystem welfare effects that result from the
deposition of total reactive nitrogen and sulfur. Because ecosystems
are diverse in biota, climate, geochemistry, and hydrology, response to
pollutant exposures can vary greatly between ecosystems. In addition,
these diverse ecosystems are not distributed evenly across the United
States. To target nitrogen and sulfur acidification and nitrogen and
sulfur enrichment, the REA addresses four main targeted ecosystem
effects on terrestrial and aquatic systems identified by the ISA (US
EPA, 2008): Aquatic acidification due to nitrogen and sulfur;
terrestrial acidification due to nitrogen and sulfur; aquatic nutrient
enrichment, including eutrophication; and terrestrial nutrient
enrichment.
In addition to these four targeted ecosystem effects, the REA also
qualitatively addresses the influence of sulfur oxides deposition on
MeHg production; nitrous oxide (N2O) effects on climate;
nitrogen effects on primary productivity and biogenic greenhouse gas
(GHG) fluxes; and phytotoxic effects on plants.
Because the targeted ecosystem effects outlined above are not
evenly distributed across the U.S., the REA identified case studies for
each targeted effects based on ecosystems identified as sensitive to
nitrogen and/or sulfur deposition effects. Eight case study areas and
two supplemental study areas (Rocky Mountain National Park and Little
Rock Lake, Wisconsin) are summarized in the REA based on ecosystem
characteristics, indicators, and ecosystem service information. Case
studies selected for aquatic acidification effects were the Adirondack
Mountains and Shenandoah National Park. Kane Experimental Forest in
Pennsylvania and Hubbard Brook Experimental Forest in New Hampshire
were selected as case studies for terrestrial acidification. Aquatic
nutrient enrichment case study locations were selected in the Potomac
River Basin upstream of Chesapeake Bay and the Neuse River Basin
upstream of the Pamlico Sound in North Carolina. The CSS communities in
southern California and the MCF communities in the San Bernardino and
Sierra Nevada Mountains of California were selected as case studies for
terrestrial nutrient enrichment. Two supplemental areas were also
chosen, one in Rocky Mountain National Park for terrestrial nutrient
enrichment and one in Little Rock Lake, Wisconsin for aquatic nutrient
enrichment.
2. Key Findings
In summary, based on case study analyses, the REA concludes that
known or anticipated adverse ecological effects are occurring under
current conditions and further concludes that these adverse effects
continue into the future. Key findings from the air quality analyses,
acidification and nutrient enrichment case studies, as well as general
conclusions from evaluating additional welfare effects, are summarized
below.

[[Page 46101]]

a. Air Quality Analyses
The air quality analyses in the REA encompass the current emissions
sources of nitrogen and sulfur, as well as atmospheric concentrations,
estimates of deposition of total nitrogen, policy-relevant background,
and non-atmospheric loadings of nitrogen and sulfur to ecosystems, both
nationwide and in the case study areas. Spatial fields of deposition
were created using wet deposition measurements from the NADP National
Trends Network and dry deposition predictions from the 2002 CMAQ model
simulation. Some key conclusions from this analysis are:
(1) Total reactive nitrogen deposition and sulfur deposition are
much greater in the East compared to most areas of the West.
(2) These regional differences in deposition correspond to the
regional differences in oxides of nitrogen and SO2
concentrations and emissions, which are also higher in the East. Oxides
of nitrogen emissions are much greater and generally more widespread
than NH3 emissions nationwide; high NH3 emissions
tend to be more local (e.g., eastern North Carolina) or sub-regional
(e.g., the upper Midwest and Plains states). The relative amounts of
oxidized versus reduced nitrogen deposition are consistent with the
relative amounts of oxides of nitrogen and NH3 emissions.
Oxidized nitrogen deposition exceeds reduced nitrogen deposition in
most of the case study areas; the major exception being the Neuse
River/Neuse River Estuary Case Study Area.
(3) Reduced nitrogen deposition exceeds oxidized nitrogen
deposition in the vicinity of local sources of NH3.
(4) There can be relatively large spatial variations in both total
reactive nitrogen deposition and sulfur deposition within a case study
area; this occurs particularly in those areas that contain or are near
a high emissions source of oxides of nitrogen, NH3 and/or
SO2.
(5) The seasonal patterns in deposition differ between the case
study areas. For the case study areas in the East, the season with the
greatest amounts of total reactive nitrogen deposition correspond to
the season with the greatest amounts of sulfur deposition. Deposition
peaks in spring in the Adirondack, Hubbard Brook Experimental Forest,
and Kane Experimental Forest case study areas, and it peaks in summer
in the Potomac River/Potomac Estuary, Shenandoah, and Neuse River/Neuse
River Estuary case study areas. For the case study areas in the West,
there is less consistency in the seasons with greatest total reactive
nitrogen and sulfur deposition in a given area. In general, both
nitrogen and/or sulfur deposition peaks in spring or summer. The
exception to this is the Sierra Nevada Range portion of the MCF Case
Study Area, in which sulfur deposition is greatest in winter.
b. Deposition-Related Aquatic Acidification
The role of aquatic acidification in two eastern United States
areas--northeastern New York's Adirondack area and the Shenandoah area
in Virginia--was analyzed in the REA to assess surface water trends in
SO4^2- and NO3^-
concentrations and ANC levels and to affirm the understanding that
reductions in deposition could influence the risk of acidification.
Monitoring data from the EPA-administered TIME)/LTM programs and the
EMAP were assessed for the years 1990 to 2006, and past, present and
future water quality levels were estimated using both steady-state and
dynamic biogeochemical models.
Although wet deposition rates for SO2 and oxides of
nitrogen in the Adirondack Case Study Area have reduced since the mid-
1990s, current concentrations are still well above pre-acidification
(1860) conditions. The MAGIC modeling predicts
NO3^- and SO4^2- are 17- and
5-fold higher today, respectively. The estimated average ANC for 44
lakes in the Adirondack Case Study Area is 62.1 [mu]eq/L (15.7 [mu]eq/L); 78 percent of all monitored lakes in the
Adirondack Case Study Area have a current risk of Elevated, Severe, or
Acute. Of the 78 percent, 31 percent experience episodic acidification,
and 18 percent are chronically acidic today.
(1) Based on the steady-state critical load model for the year
2002, 18 percent, 28 percent, 44 percent, and 58 percent of 169 modeled
lakes received combined total sulfur and nitrogen deposition that
exceeded critical loads corresponding to ANC limits of 0, 20, 50, and
100 [mu]eq/L respectively.
(2) Based on a deposition scenario that maintains current emission
levels to 2020 and 2050, the simulation forecast indicates no
improvement in water quality in the Adirondack Case Study Area. The
percentage of lakes within the Elevated to Acute Concern classes
remains the same in 2020 and 2050.
(3) Since the mid-1990s, streams in the Shenandoah Case Study Area
have shown slight declines in NO3 and
SO4^2- concentrations in surface waters. The ANC
levels increased from about 50 [mu]eq/L in the early 1990s to >75
[mu]eq/L until 2002, when ANC levels declined back to 1991-1992 levels.
Current concentrations are still above pre-acidification (1860)
conditions. The MAGIC modeling predicts surface water concentrations of
NO3 and SO4^2- are 10- and 32-fold
higher today, respectively. The estimated average ANC for 60 streams in
the Shenandoah Case Study Area is 57.9 [mu]eq/L (4.5
[mu]eq/L). Fifty-five percent of all monitored streams in the
Shenandoah Case Study Area have a current risk of Elevated, Severe, or
Acute. Of the 55 percent, 18 percent experience episodic acidification,
and 18 percent are chronically acidic today.
(4) Based on the steady-state critical load model for the year
2002, 52 percent, 72 percent, 85 percent and 93 percent of 60 modeled
streams received combined total sulfur and nitrogen deposition that
exceeded critical loads corresponding to ANC limits of 0, 20, 50, and
100 [mu]eq/L respectively.
(5) Based on a deposition scenario that maintains current emission
levels to 2020 and 2050, the simulation forecast indicates that a large
number of streams would still have Elevated to Acute problems with
acidity.
c. Deposition-Related Terrestrial Acidification
The role of terrestrial acidification was examined in the REA using
a critical load analysis for sugar maple and red spruce forests in the
eastern U.S. by using the BC/Al ratio in acidified forest soils as an
indicator to assess the impact of nitrogen and sulfur deposition on
tree health. These are the two most commonly studied species in North
America for impacts of acidification. At a BC/Al ratio of 1.2, red
spruce growth can be reduced by 20 percent. Sugar maple growth can be
reduced by 20 percent at a BC/Al ratio of 0.6. Key findings of the case
study are summarized below.
(1) Case study results suggest that the health of at least a
portion of the sugar maple and red spruce growing in the U.S. may have
been compromised with acidifying total nitrogen and sulfur deposition
in 2002. The 2002 CMAQ/NADP total nitrogen and sulfur deposition levels
exceeded three selected critical loads in 3 percent to 75 percent of
all sugar maple plots across 24 states. The three critical loads ranged
from 6,008 to 107 eq/ha/yr for the BC/Al ratios of 0.6, 1.2, and 10.0
(increasing levels of tree protection). The 2002 CMAQ/NADP total
nitrogen and sulfur deposition levels exceeded three selected critical
loads in 3 percent to 36 percent of all red spruce plots across eight
states. The three critical loads

[[Page 46102]]

ranged from 4,278 to 180 eq/ha/yr for the Bc/Al ratios of 0.6, 1.2, and
10.0 (increasing levels of tree protection).
(2) The SMB model assumptions made for base cation weathering (Bcw)
and forest soil ANC input parameters are the main sources of
uncertainty since these parameters are rarely measured and require
researchers to use default values.
(3) The pattern of case study results suggests that nitrogen and
sulfur acidifying deposition in the sugar maple and red spruce forest
areas studied were similar in magnitude to the critical loads for those
areas and both ecosystems are likely to be sensitive to any future
changes in the levels of deposition.
d. Deposition-Related Aquatic Nutrient Enrichment
The role of nitrogen deposition in two main stem rivers feeding
their respective estuaries was analyzed in the REA to determine if
decreases in deposition could influence the risk of eutrophication as
predicted using the ASSETS EI scoring system in tandem with SPARROW
modeling. This modeling approach provides a transferrable,
intermediate-level analysis of the linkages between atmospheric
deposition and receiving waters, while providing results on which
conclusions could be drawn. A summary of findings follows:
(1) The 2002 CMAQ/NADP results showed that an estimated 40,770,000
kilograms (kg) of total nitrogen was deposited in the Potomac River
watershed. The SPARROW modeling predicted that 7,380,000 kg N/yr of the
deposited nitrogen reached the estuary (20 percent of the total load to
the estuary). The overall ASSETS EI for the Potomac River and Potomac
Estuary was Bad (based on all sources of N).
(2) To improve the Potomac River and Potomac Estuary ASSETS EI
score from Bad to Poor, a decrease of at least 78 percent in the 2002
total nitrogen atmospheric deposition load to the watershed would be
required.
(3) The 2002 CMAQ/NADP results showed that an estimated 18,340,000
kg of total nitrogen was deposited in the Neuse River watershed. The
SPARROW modeling predicted that 1,150,000 kg N/yr of the deposited
nitrogen reached the estuary (26 percent of the total load to the
estuary). The overall ASSETS EI for the Neuse River/Neuse River Estuary
was Bad.
(4) It was found that the Neuse River/Neuse River Estuary ASSETS EI
score could not be improved from Bad to Poor with decreases only in the
2002 atmospheric deposition load to the watershed. Additional
reductions would be required from other nitrogen sources within the
watershed.
The small effect of decreasing atmospheric deposition in the Neuse
River watershed is because the other nitrogen sources within the
watershed are more influential than atmospheric deposition in affecting
the total nitrogen loadings to the Neuse River Estuary, as estimated
with the SPARROW model. A water body's response to nutrient loading
depends on the magnitude (e.g., agricultural sources have a higher
influence in the Neuse than in the Potomac), spatial distribution, and
other characteristics of the sources within the watershed; therefore a
reduction in nitrogen deposition does not always produce a linear
response in reduced load to the estuary, as demonstrated by these two
case studies.
e. Deposition-Related Terrestrial Nutrient Enrichment
California CSS and MCF communities were the focus of the
Terrestrial Nutrient Enrichment Case Studies of the REA. Geographic
information systems analysis supported a qualitative review of past
field research to identify ecological benchmarks associated with CSS
and mycorrhizal communities, as well as MCF's nutrient-sensitive
acidophyte lichen communities, fine-root biomass in Ponderosa pine and
leached nitrate in receiving waters. These benchmarks, ranging from 3.1
to 17 kg N/ha/yr, were compared to 2002 CMAQ/NADP data to discern any
associations between atmospheric deposition and changing communities.
Evidence supports the finding that nitrogen alters CSS and MCF. Key
findings include the following:
(1) The 2002 CMAQ/NADP nitrogen deposition data show that the 3.3
kg N/ha/yr benchmark has been exceeded in more than 93 percent of CSS
areas (654,048 ha). This suggests that such deposition is a driving
force in the degradation of CSS communities. One potentially
confounding factor is the role of fire. Although CSS decline has been
observed in the absence of fire, the contributions of deposition and
fire to the CSS decline require further research. The CSS is fragmented
into many small parcels, and the 2002 CMAQ/NADP 12-km grid data are not
fine enough to fully validate the relationship between CSS
distribution, nitrogen deposition, and fire.
(2) The 2002 CMAQ/NADP nitrogen deposition data exceeds the 3.1 kg
N/ha/yr benchmark in more than 38% (1,099,133 ha) of MCF areas, and
nitrate leaching has been observed in surface waters. Ozone effects
confound nitrogen effects on MCF acidophyte lichen, and the
interrelationship between fire and nitrogen cycling requires additional
research.
f. Additional Effects
Ecological effects have also been documented across the U.S. where
elevated nitrogen deposition has been observed, including the eastern
slope of the Rocky Mountains where shifts in dominant algal species in
alpine lakes have occurred where wet nitrogen deposition was only about
1.5 kg N/ha/yr. High alpine terrestrial communities have a low capacity
to sequester nitrogen deposition, and monitored deposition exceeding 3
to 4 kg N/ha/yr could lead to community-level changes in plant species,
lichens and mycorrhizae.
Additional welfare effects are documented, but examined less
extensively, in the REA. These effects include qualitative discussions
related to visibility and materials damage, such as corrosion, erosion,
and soiling of paint and buildings which are being addressed in the PM
NAAQS review currently underway. A discussion of the causal
relationship between sulfur deposition (as sulfate,
SO4^2-) and increased mercury methylation in
wetlands and aquatic environments is also included in the REA. On this
subject the REA concludes that decreases in SO4^2-
deposition will likely result in decreases in MeHg concentration;
however, spatial and biogeochemical variations nationally hinder
establishing large scale dose-response relationships.
Several additional issues concerning oxides of nitrogen were
addressed in the REA. Consideration was also given to N2O, a
potent GHG. The REA concluded that it is most appropriate to analyze
the role of N2O in the context of all of the GHGs rather
than as part of the REA for this review. The REA considered nitrogen
deposition and its correlation with the rate of photosynthesis and net
primary productivity. Nitrogen addition ranging from 15.4 to 300 kg N/
ha/yr is documented as increasing wetland N2O production by
an average of 207 percent across all ecosystems. Nitrogen addition
ranging from 30 to 240 kg N/ha/yr increased CH4 emissions by
115 percent, averaged across all ecosystems, and methane uptake was
reduced by 38 percent averaged across all ecosystems when nitrogen
addition ranged from 10 to 560 kg N/ha/yr, but reductions were only
significant for coniferous and deciduous forests. The heterogeneity of
ecosystems across the U.S., however, introduces variations into dose-
response relationships.

[[Page 46103]]

The phytotoxic effects of oxides of nitrogen and sulfur on
vegetation were also briefly discussed in the REA which concluded that
since a unique secondary NAAQS exists for SO2, and
concentrations of nitric oxide (NO), NO2 and PAN are rarely
high enough to have phytotoxic effects on vegetation, further
assessment was not warranted at this time.
3. Conclusions on Effects
For aquatic and terrestrial acidification effects, a similar
conceptual approach was used (critical loads) to evaluate the impacts
of multiple pollutants on an ecological endpoint, whereas the
approaches used for aquatic and terrestrial nutrient enrichment were
fundamentally distinct. Although the ecological indicators for aquatic
and terrestrial acidification (i.e., ANC and BC/Al) are very different,
both ecological indicators are well-correlated with effects such as
reduced biodiversity and growth. While aquatic acidification is clearly
the targeted effect area with the highest level of confidence, the
relationship between atmospheric deposition and an ecological indicator
is also quite strong for terrestrial acidification. The main drawback
with the understanding of terrestrial acidification is that the data
are based on laboratory responses rather than field measurements. Other
stressors that are present in the field but that are not present in the
laboratory may confound this relationship.
For nutrient enrichment effects, the REA utilized different types
of indicators for aquatic and terrestrial effects to assess both the
likelihood of adverse effects to ecosystems and the relationship
between adverse effects and atmospheric sources of oxides of nitrogen.
The ecological indicator chosen for aquatic nutrient enrichment, the
ASSETS EI, seems to be inadequate to relate atmospheric deposition to
the targeted ecological effect, likely due to the many other
confounding factors. Further, there is far less confidence associated
with the understanding of aquatic nutrient enrichment because of the
large contributions from non-atmospheric sources of nitrogen and the
influence of both oxidized and reduced forms of nitrogen, particularly
in large watersheds and coastal areas. However, a strong relationship
exists between atmospheric deposition of nitrogen and ecological
effects in high alpine lakes in the Rocky Mountains because atmospheric
deposition is the only source of nitrogen to these systems. There is
also a strong weight-of-evidence regarding the relationships between
ecological effects attributable to terrestrial nitrogen nutrient
enrichment; however, ozone and climate change may be confounding
factors. In addition, the response for other species or species in
other regions of the U.S. has not been quantified.

C. Adversity of Effects to Public Welfare

Characterizing a known or anticipated adverse effect to public
welfare is an important component of developing any secondary NAAQS.
According to the CAA, welfare effects include: ``Effects on soils,
water, crops, vegetation, manmade materials, animals, wildlife,
weather, visibility, and climate, damage to and deterioration of
property, and hazards to transportation, as well as effect on economic
values and on personal comfort and well-being, whether caused by
transformation, conversion, or combination with other air pollutants''
(CAA, Section 302(h)). While the text above lists a number of welfare
effects, these effects do not define public welfare in and of
themselves.
Although there is no specific definition of adversity to public
welfare, the paradigm of linking adversity to public welfare to
disruptions in ecosystem structure and function has been used broadly
by EPA to categorize effects of pollutants from the cellular to the
ecosystem level. An evaluation of adversity to public welfare might
consider the likelihood, type, magnitude, and spatial scale of the
effect as well as the potential for recovery and any uncertainties
relating to these considerations.
Similar concepts were used in past reviews of secondary NAAQS for
ozone and PM (relating to visibility), as well as in initial reviews of
effects from lead deposition. Because oxides of nitrogen and sulfur are
deposited from ambient sources into ecosystems where they affect
changes to organisms, populations and ecosystems, the concept of
adversity to public welfare as a result of alterations in structure and
function of ecosystems is an appropriate consideration for this review.
Based on information provided in the PA, the following section
discusses how ecological effects from deposition of oxides of nitrogen
and sulfur relate to adversity to public welfare. In the PA, public
welfare was discussed in terms of loss of ecosystem services (defined
below), which in some cases can be monetized. Each of the four main
effect areas (aquatic and terrestrial acidification and aquatic and
terrestrial nutrient over-enrichment) are discussed including current
ecological effects and associated ecosystem services.
1. Ecosystem Services
The PA defines ecosystem services as the benefits individuals and
organizations obtain from ecosystems. Ecosystem services can be
classified as provisioning (food and water), regulating (control of
climate and disease), cultural (recreational, existence, spiritual,
educational), and supporting (nutrient cycling). Conceptually, changes
in ecosystem services may be used to aid in characterizing a known or
anticipated adverse effect to public welfare. In the REA and PA
ecosystem services are discussed as a method of assessing the magnitude
and significance to the public of resources affected by ambient
concentrations of oxides of nitrogen and sulfur and deposition in
sensitive ecosystems.
The EPA has in previous NAAQS reviews defined ecological goods and
services for the purposes of a Regulatory Impact Analysis as the
``outputs of ecological functions or processes that directly or
indirectly contribute to social welfare or have the potential to do so
in the future. Some outputs may be bought and sold, but most are not
marketed.'' It is especially important to acknowledge that it is
difficult to measure and/or monetize the goods and services supplied by
ecosystems. It can be informative in characterizing adversity to public
welfare to attempt to place an economic valuation on the set of goods
and services that have been identified with respect to a change in
policy; however it must be noted that this valuation will be incomplete
and illustrative only.
Knowledge about the relationships linking ambient concentrations
and ecosystem services is considered in the PA as one method by which
to inform a policy judgment on a known or anticipated adverse public
welfare effect. For example, a change in an ecosystem structure and
process, such as foliar injury, would be classified as an ecological
effect, with the associated changes in ecosystem services, such as
primary productivity, food availability, forest products, and
aesthetics (e.g., scenic viewing), classified as public welfare
effects. Additionally, changes in biodiversity would be classified as
an ecological effect, and the associated changes in ecosystem
services--productivity, existence (nonuse) value, recreational viewing
and aesthetics--would also be classified as public welfare effects.
As described in chapters 4 and 5 of the REA, case study analyses
were performed that link deposition in sensitive ecosystems to changes
in a

[[Page 46104]]

given ecological indicator (e.g., for aquatic acidification, to changes
in ANC) and then to changes in ecosystems. Appendix 8 of the REA links
the changes in ecosystems to the services they provide (e.g., fish
species richness and its influence on recreational fishing). To the
extent possible for each targeted effect area, the REA linked ambient
concentrations of nitrogen and sulfur (i.e., ambient air quality
indicators) to deposition in sensitive ecosystems (i.e., exposure
pathways), and then to system response as measured by a given
ecological indicator (e.g., lake and stream acidification as measured
by ANC). The ecological effect (e.g., changes in fish species richness)
was then, where possible, associated with changes in ecosystem services
and the corresponding public welfare effects (e.g., recreational
fishing).
2. Effects on Ecosystem Services
The process used to link ecological indicators to ecosystem
services is discussed extensively in appendix 8 of the REA. In brief,
for each case study area assessed, the ecological indicators are linked
to an ecological response that is subsequently linked to associated
services to the extent possible. For example, in the case study for
aquatic acidification the chosen ecological indicator is ANC which can
be linked to the ecosystem service of recreational fishing. Although
recreational fishing losses are the only service effects that can be
independently quantified or monetized at this time, there are numerous
other ecosystem services that may be related to the ecological effects
of acidification.
While aquatic acidification is the focus of this proposed standard,
the other effect areas were also analyzed in the REA and these
ecosystems are being harmed by nitrogen and sulfur deposition and will
obtain some measure of protection with any decrease in that deposition
regardless of the reason for the decrease. The following summarizes the
current levels of specific ecosystem services for aquatic and
terrestrial acidification and aquatic and terrestrial nutrient over-
enrichment and attempts to quantify and when possible monetize the harm
to public welfare, as represented by ecosystem services, due to
nitrogen and sulfur deposition.
a. Aquatic Acidification
Acidification of aquatic ecosystems primarily affects the ecosystem
services that are derived from the fish and other aquatic life found in
surface waters. In the northeastern United States, the surface waters
affected by acidification are not a major source of commercially raised
or caught fish; however, they are a source of food for some
recreational and subsistence fishers and for other consumers. Although
data and models are available for examining the effects on recreational
fishing, relatively little data are available for measuring the effects
on subsistence and other consumers. Inland waters also provide
aesthetic and educational services along with non-use services, such as
existence value (protection and preservation with no expectation of
direct use). In general, inland surface waters such as lakes, rivers,
and streams also provide a number of regulating services, playing a
role in hydrological regimes and climate regulation. There is little
evidence that acidification of freshwaters in the northeastern U.S. has
significantly degraded these specific services; however, freshwater
ecosystems also provide biological control services by providing
environments that sustain delicate aquatic food chains. The toxic
effects of acidification on fish and other aquatic life impair these
services by disrupting the trophic structure of surface waters.
Although it is difficult to quantify these services and how they are
affected by acidification, it is worth noting that some of these
services may be captured through measures of provisioning and cultural
services. For example, these biological control services may serve as
``intermediate'' inputs that support the production of ``final''
recreational fishing and other cultural services.
As summarized in Chapter 4 of the PA, recent studies indicate that
acidification of lakes and streams can result in significant loss in
economic value. For example, data indicate that more than 9 percent of
adults in the northeastern part of the country participate annually in
freshwater fishing yielding 140 million freshwater fishing days. Each
fishing day has an estimated average value per day of $35. Therefore,
the implied total annual value of freshwater fishing in the
northeastern U.S. was $5 billion in 2006. Embedded in these numbers is
a degree of harm to recreational fishing services due to acidification
that has occurred over time. These harms have not been quantified on a
regional scale; however, a case study was conducted in the Adirondacks
area (US EPA, 2011, section 4.4.2).
In the Adirondacks case study, estimates of changes in recreational
fishing services were determined, as well as changes more broadly in
``cultural'' ecosystem services (including recreational, aesthetic, and
nonuse services). First, the MAGIC model (US EPA, 2009, Appendix 8 and
section 2.2) was applied to 44 lakes to predict what ANC levels would
be under both ``business as usual'' conditions (i.e., allowing for some
decline in deposition due to existing regulations) and pre-emission
(i.e., background) conditions. Second, to estimate the recreational
fishing impacts of aquatic acidification in these lakes, an existing
model of recreational fishing demand and site choice was applied. This
model predicts how recreational fishing patterns in the Adirondacks
would differ and how much higher the average annual value of
recreational fishing services would be for New York residents if lake
ANC levels corresponded to background (rather than business as usual)
conditions. To estimate impacts on a broader category of cultural (and
some provisioning) ecosystem services, results from the Banzhaf et al
(2006) valuation survey of New York residents were adapted and applied
to this context. The survey used a contingent valuation approach to
estimate the average annual household willingness to pay (WTP) for
future reductions in the percent of Adirondack lakes impaired by
acidification. The focus of the survey was on impacts on aquatic
resources. Pretesting of the survey indicated that respondents
nonetheless tended to assume that benefits would occur in the condition
of birds and forests as well as in recreational fishing.
By extrapolating the 44 lake Adirondack case study to all 3,000
Adirondack lakes and by applying the WTP survey results to all New York
residents, the study estimated aggregated benefits between $300 and
$800 million annually for the equivalent of improving lakes in the
Adirondacks region to an ANC level of 50 [mu]eq/L. The REA estimated 44
percent of the Adirondack lakes currently fall below an ANC of 50
[mu]eq/L. Several states have set goals for improving the acid status
of lakes and streams, generally targeting ANC in the range of 50 to 60
[mu]eq/L, and have engaged in costly activities to decrease
acidification.
These results imply significant value to the public in addition to
those derived from recreational fishing services. Note that the results
are only applicable to improvements in the Adirondacks valued by
residents of New York. If similar benefits exist in other acid-impacted
areas, benefits for the nation as a whole could be substantial. The
analysis provides results on only a subset of the impacts of
acidification on ecosystem services and suggests that the

[[Page 46105]]

overall impact on these services is likely to be substantial.
b. Terrestrial Acidification
Chapters 4.4.3 and 4.4.4 of the PA review several economic studies
of areas sensitive to terrestrial acidification. Forests in the
northeastern U.S. provide several important and valuable provisioning
ecosystem services, which are reflected in the production and sales of
tree products. Sugar maples are a particularly important commercial
hardwood tree species in the United States, producing timber and maple
syrup that provide hundreds of millions of dollars in economic value
annually. Red spruce is also used in a variety of wood products and
provides up to $100 million in economic value annually. Although the
data do not exist to directly link acidification damages to economic
values of lost recreational ecosystem services in forests, these
resources are valuable to the public. A recent study, reviewed in the
PA, suggests that the total annual value of recreational off-road
driving was more than $9 billion and the value of hunting and wildlife
viewing was more than $4 billion each in the northeastern States. The
EPA is not able to quantify at this time the specific effects on these
values of acid deposition, or of any specific reductions in deposition,
relative to the effects of many other factors that may affect them.
c. Nutrient Enrichment
Chapters 4.4.5 and 4.4.6 of the PA summarize economic studies of
east coast estuaries affected by nutrient over-enrichment or
eutrophication. Estuaries in the eastern United States are important
for fish and shellfish production. The estuaries are capable of
supporting large stocks of resident commercial species, and they serve
as the breeding grounds and interim habitat for several migratory
species. To provide an indication of the magnitude of provisioning
services associated with coastal fisheries, from 2005 to 2007, the
average value of total catch was $1.5 billion per year in 15 East Coast
states. Estuaries also provide an important and substantial variety of
cultural ecosystem services, including water-based recreational and
aesthetic services. For example, data indicate that 4.8 percent of the
population in coastal states from North Carolina to Massachusetts
participated in saltwater fishing, with a total of 26 million saltwater
fishing days in 2006. Based on estimates in the PA, total recreational
value from these saltwater fishing days was approximately $1.3 billion.
Recreational participation estimates for 1999-2000 showed almost 6
million individuals participated in motorboating in coastal states from
North Carolina to Massachusetts. The aggregate value of these coastal
motorboating outings was $2 billion per year. EPA is not able to
quantify at this time the specific effects on these values of nitrogen
deposition, or of any specific reductions in deposition, relative to
the effects of many other factors that may affect them.
Terrestrial ecosystems can also suffer from nutrient over-
enrichment. Each ecosystem is different in its composition of species
and nutrient requirements. Changes to individual ecosystems from
changes in nitrogen deposition can be hard to assess economically.
Relative recreational values are often determined by public use
information. Chapter 4.4.7 of the PA reviewed studies related to park
use in California. Data from California State Parks indicate that in
2002, 68.7 percent of adult residents participated in trail hiking for
an average of 24.1 days per year. The analyses in the PA indicate that
the aggregate annual benefit for California residents from trail hiking
in 2007 was $11.59 billion. EPA is not able to quantify at this time
the specific effects on these values of nitrogen deposition, or of any
specific reductions in deposition, relative to the effects of many
other factors that may affect them.
The PA also identified fire regulation as a service that could be
affected by nutrient over-enrichment of the CSS and MCF ecosystems by
encouraging growth of more flammable grasses, increasing fuel loads,
and altering the fire cycle. Over the 5-year period from 2004 to 2008,
Southern California experienced, on average, over 4,000 fires per year,
burning, on average, over 400,000 acres per year. It is not possible at
this time to quantify the contribution of nitrogen deposition, among
many other factors, to increased fire risk.
3. Summary
Adversity to public welfare can be understood by looking at how
deposition of oxides of nitrogen and sulfur affect the ecological
functions of an ecosystem (see II.A.), and then understanding the
ecosystem services that are degraded. The monetized value of the
ecosystem services provided by ecosystems that are sensitive to
deposition of oxides of nitrogen and sulfur are in the billions of
dollars each year, though it is not possible to quantify or monetize at
this time the effects on these values of nitrogen and sulfur deposition
or of any changes in deposition that may result from new secondary
standards. Many lakes and streams are known to be degraded by acidic
deposition which affects recreational fishing and tourism. Forest
growth is likely suffering from acidic deposition in sensitive areas
affecting red spruce and sugar maple timber production, sugar maple
syrup production, hiking, aesthetic enjoyment and tourism. Nitrogen
deposition contributes significantly to eutrophication in many
estuaries affecting fish production, swimming, boating, aesthetic
enjoyment and tourism. Ecosystem services are likely affected by
nutrient enrichment in many natural and scenic terrestrial areas,
affecting biodiversity, including habitat for rare and endangered
species, fire control, hiking, aesthetic enjoyment and tourism.

D. Adequacy of the Current Standards

An important issue to be addressed in the current review of the
secondary standards for oxides of nitrogen and sulfur is whether, in
view of the scientific evidence reflected in the ISA, additional
information on exposure and risk discussed in the REA, and conclusions
drawn from the PA, the existing standards provide adequate protection.
The Administrator therefore, has considered the extent to which the
current standards are adequate for the protection of public welfare.
Having reached the general conclusion that aquatic and terrestrial
ecosystems can be degraded by deposition of oxides of nitrogen and
sulfur, it is then necessary to first evaluate the appropriateness (in
terms of form and structure) of the current standards to address the
ecological effects of oxides of nitrogen and sulfur as well as the
adequacy of the current secondary standards for oxides of nitrogen and
sulfur to provide requisite protection by considering to what degree
risks to sensitive ecosystems would be expected to occur in areas that
meet the current standards. Conclusions regarding the adequacy of the
current standards are based on the available ecological effects,
exposure and risk-based evidence. In evaluating the strength of this
information, EPA has taken into account the uncertainties and
limitations in the scientific evidence. This section addresses the
adequacy of the current standards to protect against direct exposure
effects on plants from oxides of nitrogen and sulfur, the
appropriateness of the current structure of the standards to address
deposition-related effects of oxides of nitrogen and sulfur on
sensitive ecosystems and finally, the adequacy of such standards to
protect against adverse effects related to the deposition of oxides of
nitrogen and sulfur.

[[Page 46106]]

1. Adequacy of the Current Standards for Direct Effects
The current secondary oxides of nitrogen and sulfur standards are
intended to protect against adverse effects to public welfare. For
oxides of nitrogen, the current secondary standard was set identical to
the primary standard,\3\ i.e., an annual standard set for
NO2 to protect against adverse effects on vegetation from
direct exposure to ambient oxides of nitrogen. For oxides of sulfur,
the current secondary standard is a 3-hour standard intended to provide
protection for plants from the direct foliar damage associated with
atmospheric concentrations of SO2. It is appropriate to
consider whether the current standards are adequate to protect against
the direct effects on vegetation resulting from ambient NO2
and SO2 which were the basis for the current secondary
standards. The ISA concluded that there was sufficient evidence to
infer a causal relationship between exposure to SO2, NO,
NO2 and PAN and injury to vegetation. Additional research on
acute foliar injury has been limited and there is no evidence to
suggest foliar injury below the levels of the current secondary
standards for oxides of nitrogen and sulfur. There is sufficient
evidence to suggest that the levels of the current standards are likely
adequate to protect against direct phytotoxic effects.
---------------------------------------------------------------------------

\3\ The current primary NO2 standard has recently
been changed to the 3-year average of the 98th percentile of the
annual distribution of the 1 hour daily maximum of the concentration
of NO2. The current secondary standard remains as it was
set in 1971.
---------------------------------------------------------------------------

2. Appropriateness and Adequacy of the Current Standards for
Deposition-Related Effects
This section addresses two concepts necessary to evaluate the
current standards in the context of deposition related effects. First,
appropriateness of the current standards is considered with regard to
indicator, form, level and averaging time. This discussion centers
around the ability of the current standards to evaluate and provide
protection against deposition related effects that vary spatially and
temporally. It includes particular emphasis on the indicators and forms
of the current standards and the degree to which they are ecologically
relevant with regard to deposition related effects. Second, this
section evaluates the current standards in terms of adequacy of
protection.
a. Appropriateness
The ISA has established that the major effects of concern for this
review of the oxides of nitrogen and sulfur standards are associated
with deposition of nitrogen and sulfur caused by atmospheric
concentrations of oxides of nitrogen and sulfur. The current standards
are not directed toward depositional effects, and none of the elements
of the current NAAQS--indicator, form, averaging time, and level--are
suited for addressing the effects of nitrogen and sulfur deposition.
Five issues arise that call into question the ecological relevance
of the structure of the current secondary standards for oxides of
nitrogen and sulfur.
(1) The current SO2 secondary standard (0.5 ppm
SO2 over a 3-hour average) does not utilize an exposure
period that is relevant for ecosystem impacts. The majority of
deposition related impacts are associated with depositional loads that
occur over periods of months to years. This differs significantly from
exposures associated with hourly concentrations of SO2 as
measured by the current secondary standard. By addressing short-term
concentrations, the current SO2 secondary standard, while
protective against direct foliar effects from gaseous oxides of sulfur,
does not take into account the findings of effects in the ISA, which
notes the relationship between annual deposition of sulfur and
acidification effects which are likely to be more severe and widespread
than phytotoxic effects under current ambient conditions, and include
effects from long term deposition as well as short term. Acidification
is a process that occurs over time because the ability of an aquatic
system to counteract acidic inputs is reduced as natural buffers are
used more rapidly than they can be replaced through geologic
weathering. The relevant period of exposure for ecosystems is,
therefore, not the exposures captured in the short averaging time of
the current SO2 secondary standard. The current secondary
standard for oxides of nitrogen is an annual standard (0.053 ppm
averaged over 1 year) and as such is more ecologically relevant.
(2) Current standards do not utilize appropriate atmospheric
indicators. Nitrogen dioxide and SO2 are used as the
component of oxides of nitrogen and sulfur that are measured, but they
do not provide a complete link to the direct effects on ecosystems from
deposition of oxides of nitrogen and sulfur as they do not capture all
relevant chemical species of oxidized nitrogen and oxidized sulfur that
contribute to deposition. The ISA provides evidence that deposition
related effects are linked with total nitrogen and total sulfur
deposition, and thus all forms of oxidized nitrogen and oxidized sulfur
that are deposited will contribute to effects on ecosystems. Thus, by
using atmospheric NO2 and SO2 concentrations as
indicators, the current standards address only a fraction of total
atmospheric oxides of nitrogen and sulfur, and do not take into account
the effects from deposition of total atmospheric oxides of nitrogen and
sulfur. This suggests that more comprehensive atmospheric indicators
should be considered in designing ecologically relevant standards.
(3) Current standards reflect separate assessments of the two
individual pollutants, NO2 and SO2, rather than
assessing the joint impacts of deposition to ecosystems. Recognizing
the role that each pollutant plays in jointly affecting ecosystem
indicators, functions, and services is vital to developing a meaningful
standard. The clearest example of this interaction is in assessment of
the impacts of acidifying deposition on aquatic ecosystems.
Acidification in an aquatic ecosystem depends on the total acidifying
potential of the deposition of both nitrogen and sulfur from both
atmospheric deposition of oxides of nitrogen and sulfur as well as the
inputs from other sources of nitrogen and sulfur such as reduced
nitrogen and non-atmospheric sources. It is the joint impact of the two
pollutants that determines the ultimate effect on organisms within the
ecosystem, and critical ecosystem functions such as habitat provision
and biodiversity. Standards that are set independently are less able to
account for the contribution of the other pollutant. This suggests that
interactions between oxides of nitrogen and oxides of sulfur should be
a critical element of the conceptual framework for ecologically
relevant standards. There are also important interactions between
oxides of nitrogen and sulfur and reduced forms of nitrogen, which also
contribute to acidification and nutrient enrichment. It is important
that the structure of the standards address the role of reduced
nitrogen in determining the ecological effects resulting from
deposition of atmospheric oxides of nitrogen and sulfur. Consideration
will also have to be given to total loadings as ecosystems respond to
all sources of nitrogen and sulfur.
(4) Current standards do not take into account variability in
ecosystem sensitivity. Ecosystems are not uniformly distributed either
spatially or temporally in their sensitivity to oxides of nitrogen and
sulfur. Therefore, failure to account for the major determinants of
variability, including geological and soil

[[Page 46107]]

characteristics related to the sensitivity to acidification or nutrient
enrichment as well as atmospheric and landscape characteristics that
govern rates of deposition, may lead to standards that do not provide
requisite levels of protection across ecosystems. The current
structures of the standards do not address the complexities in the
responses of ecosystems to deposition of oxides of nitrogen and sulfur.
Ecosystems contain complex groupings of organisms that respond in
various ways to the alterations of soil and water that result from
deposition of nitrogen and sulfur compounds. Different ecosystems
therefore respond in different ways depending on a multitude of factors
that control how deposition is integrated into the system. For example,
the same levels of deposition falling on limestone dominated soils have
a very different effect from those falling on shallow glaciated soils
underlain with granite. One system may over time display no obvious
detriment while the other may experience a catastrophic loss in fish
communities. This degree of sensitivity is a function of many
atmospheric factors that control rates of deposition as well as
ecological factors that control how an ecosystem responds to that
deposition. The current standards do not take into account spatial and
seasonal variations, not only in depositional loadings, but also in
sensitivity of ecosystems exposed to those loadings. Based on the
discussion summarized above, the PA concludes that the current
secondary standards for oxides of nitrogen and oxides of sulfur are not
ecologically relevant in terms of averaging time, form, level or
indicator.
b. Adequacy of Protection
As described in the PA, ambient conditions in 2005 indicate that
the current SO2 and NO2 secondary standards were
not exceeded at that time (US EPA, 2011, Figures 6-1 and 6-2) in
locations where negative ecological effects have been observed. In many
locations, SO2 and NO2 concentrations are
substantially below the levels of the secondary standards. This pattern
suggests that levels of deposition and any negative effects on
ecosystems due to deposition of oxides of nitrogen and sulfur under
recent conditions are occurring even though areas meet or are below
current standards. In addition, based on conclusions in the REA, these
levels will not decline in the future to levels below which it is
reasonable to anticipate effects.
In determining the adequacy of the current secondary standards for
oxides of nitrogen and sulfur the PA considered the extent to which
ambient deposition contributes to loadings in ecosystems. Since the
last review of the secondary standard for oxides of nitrogen, a great
deal of information on the contribution of atmospheric deposition
associated with ambient oxides of nitrogen has become available. The
REA presents a thorough assessment of the contribution of oxidized
nitrogen to nitrogen deposition throughout the U.S., and the relative
contributions of ambient oxidized and reduced forms of nitrogen. The
REA concludes that based on that analysis, ambient oxides of nitrogen
are a significant component of atmospheric nitrogen deposition, even in
areas with relatively high rates of deposition of reduced nitrogen. In
addition, atmospheric deposition of oxidized nitrogen contributes
significantly to total nitrogen loadings in nitrogen sensitive
ecosystems.
The ISA summarizes the available studies of relative nitrogen
contribution and finds that in much of the U.S., oxides of nitrogen
contribute from 50 to 75 percent of total atmospheric deposition
relative to total reactive nitrogen, which includes oxidized and
reduced nitrogen species (US EPA, 2008, section 2.8.4). Although the
proportion of total nitrogen loadings associated with atmospheric
deposition of nitrogen varies across locations, the ISA indicates that
atmospheric nitrogen deposition is the main source of new anthropogenic
nitrogen to most headwater streams, high elevation lakes, and low-order
streams. Atmospheric nitrogen deposition contributes to the total
nitrogen load in terrestrial, wetland, freshwater and estuarine
ecosystems that receive nitrogen through multiple pathways. In several
large estuarine systems, including the Chesapeake Bay, atmospheric
deposition accounts for between 10 and 40 percent of total nitrogen
loadings (US EPA, 2008).
Atmospheric concentrations of oxides of sulfur account for nearly
all sulfur deposition in the US. For the period 2004-2006, mean sulfur
deposition in the U.S. was greatest east of the Mississippi River with
the highest deposition amount, 21.3 kg S/ha-yr, in the Ohio River
Valley where most recording stations reported 3-year averages >10 kg S/
ha-yr. Numerous other stations in the East reported S deposition >5 kg
S/ha-yr. Total sulfur deposition in the U.S. west of the 100th meridian
was relatively low, with all recording stations reporting <2 kg S/ha-yr
and many reporting <1 kg S/ha-yr. Sulfur was primarily deposited in the
form of wet SO4^2- followed in decreasing order by
a smaller proportion of dry SO2 and a much smaller
proportion of deposition as dry SO4^2-.
As discussed throughout the REA (US EPA, 2009 and section II.B
above), there are several key areas of risk that are associated with
ambient concentrations of oxides of nitrogen and sulfur. As noted
earlier, in previous reviews of the secondary standards for oxides of
nitrogen and sulfur, the standards were designed to protect against
direct exposure of plants to ambient concentrations of the pollutants.
A significant shift in understanding of the effects of oxides of
nitrogen and sulfur has occurred since the last reviews, reflecting the
large amount of research that has been conducted on the effects of
deposition of nitrogen and sulfur to ecosystems. The most significant
current risks of adverse effects to public welfare are those related to
deposition of oxides of nitrogen and sulfur to both terrestrial and
aquatic ecosystems. These risks fall into two categories, acidification
and nutrient enrichment, which were emphasized in the REA as most
relevant to evaluating the adequacy of the existing standards in
protecting public welfare from adverse ecological effects.
i. Aquatic Acidification
The focus of the REA case studies was on determining whether
deposition of sulfur and oxidized nitrogen in locations where ambient
oxides of nitrogen and sulfur were at or below the current standards
was resulting in acidification and related effects, including episodic
acidification and mercury methylation. Based on the case studies
conducted for lakes in the Adirondacks and streams in Shenandoah
National Park (case studies are discussed more fully in section II.B
and US EPA, 2009), there is significant risk to acid sensitive aquatic
ecosystems at atmospheric concentrations of oxides of nitrogen and
sulfur at or below the current standards. The REA also supports
strongly a relationship between atmospheric deposition of oxides of
nitrogen and sulfur and loss of ANC in sensitive ecosystems and
indicates that ANC is an excellent indicator of aquatic acidification.
The REA also concludes that at levels of deposition associated with
oxides of nitrogen and sulfur concentrations at or below the current
standards, ANC levels are expected to be below benchmark values that
are associated with significant losses in fish species richness.
Significant portions of the U.S. are acid sensitive, and current
deposition levels exceed those that would allow

[[Page 46108]]

recovery of the most acid sensitive lakes in the Adirondacks (US EPA,
2008, Executive Summary). In addition, because of past loadings, areas
of the Shenandoah are sensitive to current deposition levels (US EPA,
2008, Executive Summary). Parts of the West are naturally less
sensitive to acidification and subjected to lower deposition
(particularly SOX) levels relative to the eastern United
States, and as such, less focus in the ISA is placed on the adequacy of
the existing standards in these areas, with the exception of the
mountainous areas of the West, which experience episodic acidification
due to deposition.
In describing the effects of acidification in the two case study
areas the REA uses the approach of describing benchmarks in terms of
ANC values. Many locations in sensitive areas of the U.S. have ANC
levels below benchmark levels for ANC classified as severe, elevated,
or moderate concern (US EPA, 2011, Figure 2-1). The average current ANC
levels across 44 lakes in the Adirondack case study area is 62.1
[mu]eq/L (moderate concern). However, 44 percent of lakes had
deposition levels exceeding the critical load for an ANC of 50 [mu]eq/L
(elevated), and 28 percent of lakes had deposition levels exceeding the
(higher) critical load for an ANC of 20 [mu]eq/L (severe) (US EPA,
2009, section 4.2.4.2). This information indicates that almost half of
the 44 lakes in the Adirondacks case study area are at an elevated
concern level, and almost a third are at a severe concern level. These
levels are associated with greatly diminished fish species diversity,
and losses in the health and reproductive capacity of remaining
populations. Based on assessments of the relationship between number of
fish species and ANC level in both the Adirondacks and Shenandoah
areas, the number of fish species is decreased by over half at an ANC
level of 20 [mu]eq/L relative to an ANC level at 100 [mu]eq/L (US EPA,
2009, Figure 4.2-1). When extrapolated to the full population of lakes
in the Adirondacks area using weights based on the EMAP probability
survey (US EPA, 2009, section 4.2.6.1), 36 percent of lakes exceeded
the critical load for an ANC of 50 [mu]eq/L and 13 percent of lakes
exceeded the critical load for an ANC of 20 [mu]eq/L.
Many streams in the Shenandoah case study area also have levels of
deposition that are associated with ANC levels classified as severe,
elevated, or moderate concern. The average ANC under recent conditions
for the 60 streams evaluated in the Shenandoah case study area is 57.9
[mu]eq/L, indicating moderate concern. However, 85 percent of these
streams had recent deposition exceeding the critical load for an ANC of
50 [mu]eq/L, and 72 percent exceeded the critical load for an ANC of 20
[mu]eq/L. As with the Adirondacks area, this information suggests that
ANC levels may decline in the future and significant numbers of
sensitive streams in the Shenandoah area are at risk of adverse impacts
on fish populations if recent conditions persist. Many other streams in
the Shenandoah area are also likely to experience conditions of
elevated to severe concern based on the prevalence in the area of
bedrock geology associated with increased sensitivity to acidification
suggesting that effects due to stream acidification could be widespread
in the Shenandoah area (US EPA, 2009, section 4.2.6.2).
In addition to these chronic acidification effects, the ISA notes
that ``consideration of episodic acidification greatly increases the
extent and degree of estimated effects for acidifying deposition on
surface waters'' (US EPA, 2008, section 3.2.1.6). Some studies show
that the number of lakes that could be classified as acid-impacted
based on episodic acidification is 2 to 3 times the number of lakes
classified as acid-impacted based on chronic ANC. These episodic
acidification events can have long term effects on fish populations (US
EPA, 2008, section 3.2.1.6). Under recent conditions, episodic
acidification has been observed in locations in the eastern U.S. and in
the mountainous western U.S. (US EPA, 2008, section 3.2.1.6).
The ISA, REA and PA all conclude that the current standards are not
adequate to protect against the adverse impacts of aquatic
acidification on sensitive ecosystems. A recent survey, as reported in
the ISA, found sensitive streams in many locations in the U.S.,
including the Appalachian Mountains, the Coastal Plain, and the
Mountainous West (US EPA, 2008, section 4.2.2.3). In these sensitive
areas, between 1 and 6 percent of stream kilometers are chronically
acidified. The REA further concludes that both the Adirondack and
Shenandoah case study areas are currently receiving deposition from
ambient oxides of nitrogen and sulfur in excess of their ability to
neutralize such inputs. In addition, based on the current emission
scenarios, forecast modeling out to the year 2020 as well as 2050
indicates a large number of streams in these areas will still be
adversely impacted (section II.B). Based on these considerations, the
PA concludes that the current secondary NAAQS for oxides of nitrogen
and sulfur do not provide adequate protection of sensitive ecosystems
with regard to aquatic acidification.
ii. Terrestrial Acidification
Based on the terrestrial acidification case studies, Kane
Experimental Forest in Pennsylvania and Hubbard Brook Experimental
Forest described in section II.B) of sugar maple and red spruce
habitat, the REA concludes that there is significant risk to sensitive
terrestrial ecosystems from acidification at atmospheric concentrations
of NO2 and SO2 at or below the current standards.
The ecological indicator selected for terrestrial acidification is the
BC/Al, which has been linked to tree health and growth. The results of
the REA strongly support a relationship between atmospheric deposition
of oxides of nitrogen and sulfur and BC/Al, and that BC/Al is a good
indicator of terrestrial acidification. At levels of deposition
associated with oxides of nitrogen and sulfur concentrations at or
below the current standards, BC/Al levels are expected to be below
benchmark values that are associated with significant effects on tree
health and growth. Such degradation of terrestrial ecosystems could
affect ecosystem services such as habitat provisioning, endangered
species, goods production (timber, syrup, etc.) among others.
Many locations in sensitive areas of the U.S. have BC/Al levels
below benchmark levels classified as providing low to intermediate
levels of protection to tree health. At a BC/Al ratio of 1.2
(intermediate level of protection), red spruce growth can be reduced by
20 percent. At a BC/Al ratio of 0.6 (low level of protection), sugar
maple growth can be decreased by 20 percent. The REA did not evaluate
broad sensitive regions. However, in the sugar maple case study area
(Kane Experimental Forest), recent deposition levels are associated
with a BC/Al ratio below 1.2, indicating between intermediate and low
level of protection, which would indicate the potential for a greater
than 20 percent reduction in growth. In the red spruce case study area
(Hubbard Brook Experimental Forest), recent deposition levels are
associated with a BC/Al ratio slightly above 1.2, indicating slightly
better than an intermediate level of protection (US EPA, 2009, section
4.3.5.1).
Over the full range of sugar maple, 12 percent of evaluated forest
plots exceeded the critical loads for a BC/Al ratio of 1.2, and 3
percent exceeded the critical load for a BC/Al ratio of 0.6. However,
there was large variability across states. In New Jersey, 67 percent of
plots exceeded the critical load for a

[[Page 46109]]

BC/Al ratio of 1.2, while in several states on the outskirts of the
range for sugar maple (e.g. Arkansas, Illinois) no plots exceeded the
critical load for a BC/Al ratio of 1.2. For red spruce, overall 5
percent of plots exceeded the critical load for a BC/Al ratio of 1.2,
and 3 percent exceeded the critical load for a BC/Al ratio of 0.6. In
the major red spruce producing states (Maine, New Hampshire, and
Vermont), critical loads for a BC/Al ratio of 1.2 were exceeded in 0.5,
38, and 6 percent of plots, respectively.
The ISA, REA and PA all conclude that the current standards are not
adequate to protect against the adverse impacts of terrestrial
acidification on sensitive ecosystems. As stated in the REA and PA, the
main drawback, with the understanding of terrestrial acidification lies
in the sparseness of available data by which we can predict critical
loads and that the data are based on laboratory responses rather than
field measurements. Other stressors that are present in the field but
that are not present in the laboratory may confound this relationship.
The REA does however, conclude that the case study results, when
extended to a 27 state region, show that nitrogen and sulfur acidifying
deposition in the sugar maple and red spruce forest areas caused the
calculated Bc/Al ratio to fall below 1.2 (the intermediate level of
protection) in 12 percent of the sugar maple plots and 5 percent of the
red spruce plots; however, results from individual states ranged from 0
to 67 percent of the plots for sugar maple and 0 to 100 percent of the
plots for red spruce.
iii. Terrestrial Nutrient Enrichment
Nutrient enrichment effects are due to nitrogen loadings from both
atmospheric and non-atmospheric sources. Evaluation of nutrient
enrichment effects requires an understanding that nutrient inputs are
essential to ecosystem health and that specific long term levels of
nutrients in a system affect the types of species that occur over long
periods of time. Short term additions of nutrients can affect species
competition, and even small additions of nitrogen in areas that are
traditionally nutrient poor can have significant impacts on
productivity as well as species composition. Most ecosystems in the
U.S. are nitrogen-limited, so regional decreases in emissions and
deposition of airborne nitrogen compounds could lead to some decrease
in growth of the vegetation that surrounds the targeted aquatic system
but as discussed below evidence for this is mixed. Whether these
changes in plant growth are seen as beneficial or adverse will depend
on the nature of the ecosystem being assessed.
Information on the effects of changes in nitrogen deposition on
forestlands and other terrestrial ecosystems is very limited. The
multiplicity of factors affecting forests, including other potential
stressors such as ozone, and limiting factors such as moisture and
other nutrients, confound assessments of marginal changes in any one
stressor or nutrient in forest ecosystems. The ISA notes that only a
fraction of the deposited nitrogen is taken up by the forests, most of
the nitrogen is retained in the soils (US EPA, 2008, section 3.3.2.1).
In addition, the ISA indicates that forest management practices can
significantly affect the nitrogen cycling within a forest ecosystem,
and as such, the response of managed forests to nitrogen deposition
will be variable depending on the forest management practices employed
in a given forest ecosystem (US EPA, 2008, Annex C C.6.3). Increases in
the availability of nitrogen in nitrogen-limited forests via
atmospheric deposition could increase forest production over large non-
managed areas, but the evidence is mixed, with some studies showing
increased production and other showing little effect on wood production
(US EPA, 2008, section 3.3.9). Because leaching of nitrate can promote
cation losses, which in some cases create nutrient imbalances, slower
growth and lessened disease and freezing tolerances for forest trees,
the net effect of increased N on forests in the U.S. is uncertain (US
EPA, 2008, section 3.3.9).
The scientific literature has many examples of the deleterious
effects caused by excessive nitrogen loadings to terrestrial systems.
Several studies have set benchmark values for levels of N deposition at
which scientifically adverse effects are known to occur. Large areas of
the country appear to be experiencing deposition above these
benchmarks. The ISA indicates studies that have found that at 3.1 kg N/
ha/yr, the community of lichens begins to change from acidophytic to
tolerant species; at 5.2 kg N/ha/yr, the typical dominance by
acidophytic species no longer occurs; and at 10.2 kg N/ha/yr,
acidophytic lichens are totally lost from the community. Additional
studies in the Colorado Front Range of the Rocky Mountain National Park
support these findings. These three values (3.1, 5.2, and 10.2 kg/ha/
yr) are one set of ecologically meaningful benchmarks for the mixed
conifer forest (MCF) of the pacific coast regions. Nearly all of the
known sensitive communities receive total nitrogen deposition levels
above the 3.1 N kg/ha/yr ecological benchmark according to the 12 km,
2002 CMAQ/NADP data, with the exception of the easternmost Sierra
Nevadas. The MCFs in the southern portion of the Sierra Nevada forests
and nearly all MCF communities in the San Bernardino forests receive
total nitrogen deposition levels above the 5.2 N kg/ha/yr ecological
benchmark.
Coastal Sage Scrub communities are also known to be sensitive to
community shifts caused by excess nitrogen loadings. Studies have
investigated the amount of nitrogen utilized by healthy and degraded
CSS systems. In healthy stands, the authors estimated that 3.3 kg N/ha/
yr was used for CSS plant growth. It is assumed that 3.3 kg N/ha/yr is
near the point where nitrogen is no longer limiting in the CSS
community and above which level community changes occur, including
dominance by invasive species and loss of coastal sage scrub.
Therefore, this amount can be considered an ecological benchmark for
the CSS community. The majority of the known CSS range is currently
receiving deposition in excess of this benchmark. Thus, the REA
concludes that recent conditions where oxides of nitrogen ambient
concentrations are at or below the current oxides of nitrogen secondary
standards are not adequate to protect against anticipated adverse
impacts from N nutrient enrichment in sensitive ecosystems.
iv. Aquatic Nutrient Enrichment
The REA aquatic nutrient enrichment case studies focused on coastal
estuaries and revealed that while current ambient loadings of
atmospheric oxides of nitrogen are contributing to the overall
depositional loading of coastal estuaries, other non-atmospheric
sources are contributing in far greater amounts in total, although
atmospheric contributions are as large as some other individual source
types. The ability of current data and models to characterize the
incremental adverse impacts of nitrogen deposition is limited, both by
the available ecological indicators, and by the inability to attribute
specific effects to atmospheric sources of nitrogen. The REA case
studies used ASSETS EI as the ecological indicator for aquatic nutrient
enrichment. This index is a six level index characterizing overall
eutrophication risk in a water body. This indicator is not sensitive to
changes in nitrogen deposition within a single level of the index. In
addition, this type of indicator does not reflect the impact of
nitrogen deposition in conjunction with other sources of nitrogen.

[[Page 46110]]

Based on the above considerations, the REA concludes that the
ASSETS EI is not an appropriate ecological indicator for estuarine
aquatic eutrophication and that additional analysis is required to
develop an appropriate indicator for determining the appropriate levels
of protection from N nutrient enrichment effects in estuaries related
to deposition of oxides of nitrogen. As a result, EPA is unable to make
a determination as to the adequacy of the existing secondary oxides of
nitrogen standard in protecting public welfare from nitrogen nutrient
enrichment effects in estuarine aquatic ecosystems.
Additionally, nitrogen deposition can alter species composition and
cause eutrophication in freshwater systems. In the Rocky Mountains, for
example, deposition loads of 1.5 to 2 kg/ha/yr which are well within
current ambient levels are known to cause changes in species
composition in diatom communities indicating impaired water quality (US
EPA, 2008, section 3.3.5.3). This suggests that the existing secondary
standard for oxides of nitrogen does not protect such ecosystems and
their resulting services from impairment.
v. Other Effects
An important consideration in looking at the effects of deposition
of oxides of sulfur in aquatic ecosystems is the potential for
production of MeHg, a neurotoxic contaminant. The production of
meaningful amounts of MeHg requires the presence of
SO4^2- and mercury, and where mercury is present,
increased availability of SO4^2- results in
increased production of MeHg. There is increasing evidence on the
relationship between sulfur deposition and increased methylation of
mercury in aquatic environments; this effect occurs only where other
factors are present at levels within a range to allow methylation. The
production of MeHg requires the presence of SO4^2-
and mercury, but the amount of MeHg produced varies with oxygen
content, temperature, pH and supply of labile organic carbon (US EPA,
2008, section 3.4). In watersheds where changes in sulfate deposition
did not produce an effect, one or several of those interacting factors
were not in the range required for meaningful methylation to occur (US
EPA, 2008, section 3.4). Watersheds with conditions known to be
conducive to mercury methylation can be found in the northeastern
United States and southeastern Canada (US EPA, 2009, section 6).
With respect to sulfur deposition and mercury methylation, the
final ISA determined that ''[t]he evidence is sufficient to infer a
causal relationship between sulfur deposition and increased mercury
methylation in wetlands and aquatic environments.'' However, EPA did
not conduct a quantitative assessment of the risks associated with
increased mercury methylation under current conditions. As such, EPA is
unable to make a determination as to the adequacy of the existing
SO2 secondary standards in protecting against welfare
effects associated with increased mercury methylation.
vi. Summary of Adequacy Considerations
In summary, the PA concludes that currently available scientific
evidence and assessments clearly call into question the adequacy of the
current standards with regard to deposition-related effects on
sensitive aquatic and terrestrial ecosystems, including acidification
and nutrient enrichment. Further, the PA recognizes that the elements
of the current standards--indicator, averaging time, level and form--
are not ecologically relevant, and are thus not appropriate for
standards designed to provide such protection. Thus, the PA concludes
that consideration should be given to establishing a new ecologically
relevant multi-pollutant, multimedia standard to provide appropriate
protection from deposition-related ecological effects of oxides of
nitrogen and sulfur on sensitive ecosystems, with a focus on protecting
against adverse effects associated with acidifying deposition in
sensitive aquatic ecosystems.
3. CASAC Views
In a letter to the Administrator (Russell and Samet 2011a), the
CASAC Oxides of Nitrogen and Oxides of Sulfur Panel, with full
endorsement of the chartered CASAC, unanimously concluded that:

EPA staff has demonstrated through the Integrated Science
Assessment (ISA), Risk and Exposure Characterization (REA) and the
draft PA that ambient NOX and SOX can have,
and are having, adverse environmental impacts. The Panel views that
the current NOX and SOX secondary standards
should be retained to protect against direct adverse impacts to
vegetation from exposure to gas phase exposures of these two
families of air pollutants. Further, the ISA, REA and draft PA
demonstrate that adverse impacts to aquatic ecosystems are also
occurring due to deposition of NOX and SOX.
Those impacts include acidification and undesirable levels of
nutrient enrichment in some aquatic ecosystems. The levels of the
current NOX and SOX secondary NAAQS are not
sufficient, nor the forms of those standards appropriate, to protect
against adverse depositional effects; thus a revised NAAQS is
warranted.

In addition, with regard to the joint consideration of both oxides
of nitrogen and oxides of sulfur as well as the consideration of
deposition related effects, CASAC concluded that the PA had developed a
credible methodology for considering such effects. The Panel stated
that ``the Policy Assessment develops a framework for a multi-
pollutant, multimedia standard that is ecologically relevant and
reflects the combined impacts of these two pollutants as they deposit
to sensitive aquatic ecosystems.''
4. Administrator's Proposed Conclusions Concerning Adequacy of Current
Standard
Based on the above considerations and taking into account CASAC
advice, the Administrator recognizes that the purpose of the secondary
standard is to protect against ``adverse'' effects resulting from
exposure to oxides of nitrogen and sulfur, discussed above in section
II.A. The Administrator also recognizes the need for conclusions as to
the adequacy of the current standards for both direct and deposition
related effects as well as conclusions as to the appropriateness and
ecological relevance of the current standards.
In considering what constitutes an ecological effect that is also
adverse to the public welfare, the Administrator took into account the
ISA conclusions regarding the nature and strength of the effects
evidence, the risk and exposure assessment results, the degree to which
the associated uncertainties should be considered in interpreting the
results, the conclusions presented in the PA, and the views of CASAC
and members of the public. On these bases, the Administrator concludes
that the current secondary standards are adequate to protect against
direct phytotoxic effects on vegetation. Thus, the Administrator
proposes to retain the current secondary standard for oxides of
nitrogen at 53 ppb,\4\ annual average concentration, measured in the
ambient air as NO2, and the current secondary standard for
oxides of sulfur at 0.5 ppm,

[[Page 46111]]

3-hour average concentration, measured in the ambient air as
SO2.
---------------------------------------------------------------------------

\4\ The annual secondary standard for oxides of nitrogen is
being specified in units of ppb to conform to the current version of
the annual primary standard, as specified in the final rule for the
most recent review of the NO2 primary NAAQS (75 FR 6531;
February 9, 2010).
---------------------------------------------------------------------------

With regard to deposition-related effects, the Administrator has
first to consider the appropriateness of the structure of the current
standards to address ecological effects of concern. Based on the
evidence as well as considering the advice given by CASAC on this
matter, the Administrator concludes that the elements of the current
standards are not ecologically relevant and thus are not appropriate to
provide protection of ecosystems. On the subject of adequacy of
protection with regard to deposition-related effects, the Administrator
considered the full nature of ecological effects related to the
deposition of ambient oxides of nitrogen and sulfur into sensitive
ecosystems across the country. Her conclusions are based on the
evidence presented in the ISA with regard to acidification and nutrient
enrichment effects, the findings of the REA with regard to scope and
severity of the current and likely future effects of deposition, the
synthesis of both the scientific evidence and risk and exposure results
in the PA as to the adequacy of the current standards, and the advice
of both CASAC and the public. After such consideration, the
Administrator concludes that current levels of oxides of nitrogen and
sulfur are sufficient to cause acidification of both aquatic and
terrestrial ecosystems, nutrient enrichment of terrestrial ecosystems
and contribute to nutrient enrichment effects in estuaries that could
be considered adverse, and the current secondary standards do not
provide adequate protection from such effects.
Having reached these conclusions, the Administrator determines that
it is appropriate to consider alternative standards that are
ecologically relevant. These considerations support the conclusion that
the current secondary standards is neither appropriate nor adequate to
protect against deposition related effects. The Administrator's
consideration of such alternative standards is discussed below in
Section III.

III. Rationale for Proposed Decision on Alternative Multi-Pollutant
Approach to Secondary Standards for Aquatic Acidification

Having reached the conclusion that the current NO2 and
SO2 secondary standards are not adequate to provide
appropriate protection against deposition-related effects associated
with oxides of nitrogen and sulfur, the Administrator then considered
what new multi-pollutant standard might be appropriate, at this time,
to address such effects on public welfare. The Administrator recognizes
that the inherently complex and variable linkages between ambient
concentrations of nitrogen and sulfur oxides, the related deposited
forms of nitrogen and sulfur, and the ecological responses that are
associated with public welfare effects call for consideration of an
ecologically relevant design of a standard that reflects these
linkages. The Administrator also recognizes that characterization of
such complex and variable linkages will necessarily require
consideration of information and analyses that have important
limitations and uncertainties.
Despite its complexity, an ecologically relevant multi-pollutant
standard to address deposition-related effects could still
appropriately be defined in terms of the same basic elements that are
used to define any NAAQS--indicator, form, averaging time, and level.
The form would incorporate additional structural elements that reflect
relevant multi-pollutant and multimedia attributes. These structural
elements include the use of an ecological indicator, tied to the
ecological effect we are focused on, and other elements that account
for ecologically relevant factors other than ambient air
concentrations. All of these elements would be needed to enable a
linkage from ambient air indicators to the ecological indicator to
define an ecologically relevant standard. As a result, such a standard
would necessarily be more complex than the NAAQS that have been set
historically to address effects associated with ambient concentrations
of a single pollutant.
More specifically, the Administrator considered an ecologically
relevant multi-pollutant standard to address effects associated with
acidifying deposition related to ambient concentrations of oxides of
nitrogen and sulfur in sensitive aquatic ecosystems. This focus is
consistent with the information presented in the ISA, REA, and PA,
which highlighted the sufficiency of the quantity and quality of the
available evidence and assessments associated with aquatic
acidification relative to the information and assessments available for
other deposition-related effects, including terrestrial acidification
and aquatic and terrestrial nutrient enrichment. Based on its review of
these documents, CASAC agreed that aquatic acidification should be the
focus for developing a new multi-pollutant standard in this review. In
reaching conclusions about an air quality standard designed to address
deposition-related aquatic acidification effects, the Administrator
also recognizes that such a standard may also provide some degree of
protection against other deposition-related effects.
As discussed in chapter 7 of the PA, the development of a new
multi-pollutant standard to address deposition-related aquatic
acidification effects recognizes the need for consideration of a
nationally applicable standard for protection against adverse effects
of aquatic acidification on public welfare, while recognizing the
complex and heterogeneous interactions between ambient air
concentrations of nitrogen and sulfur oxides, the related deposition of
nitrogen and sulfur, and associated ecological responses. The
development of such a standard also needs to take into account the
limitations and uncertainties in the available information and analyses
upon which characterization of such interactions are based. The
approach used in the PA also recognizes that while such a standard
would be national in scope and coverage, the effects to public welfare
from aquatic acidification will not occur to the same extent in all
locations in the U.S., given the inherent variability of the responses
of aquatic systems to the effects of acidifying deposition.
As discussed above in section II, many locations in the U.S. are
naturally protected against acid deposition due to underlying
geological conditions. Likewise, some locations in the U.S., including
lands managed for commercial agriculture and forestry, are not likely
to be negatively impacted by current levels of nitrogen and sulfur
deposition. As a result, while a new ecologically relevant secondary
standard would apply everywhere, it would be structured to account for
differences in the sensitivity of ecosystems across the country. This
would allow for appropriate protection of sensitive aquatic ecosystems,
which are relatively pristine and wild and generally in rural areas,
and the services provided by such sensitive ecosystems, without
requiring more protection than is needed elsewhere.
As discussed below, the multi-pollutant standard developed in the
PA would employ (1) total reactive oxidized nitrogen (NOy)
and SOX as the atmospheric ambient air indicators; (2) a
form that takes into account variable factors, such as atmospheric and
ecosystem conditions that modify the amounts of deposited nitrogen and
sulfur; the distinction between oxidized and reduced forms of nitrogen;
effects of deposited nitrogen and sulfur on aquatic ecosystems in terms
of the ecological indicator ANC; and the

[[Page 46112]]

representativeness of water bodies within a defined spatial area; (3) a
multi-year averaging time, and (4) a standard level defined in terms of
a single, national target ANC value that, in the context of the above
form, identifies the levels of concentrations of NOy and
SOX in the ambient air that would meet the standard. The
form of such a standard has been defined by an index, AAI, which
reflects the relationship between ambient concentrations of
NOy and SOX and aquatic acidification effects
that result from nitrogen and sulfur deposition related to these
ambient concentrations.
In presenting the considerations associated with such an air
quality standard to address deposition-related aquatic acidification
effects, the following sections focus on each element of the standard,
including indicator (section III.A), form (section III.B), averaging
time (section III.C), and level (section III.D). Alternative
combinations of levels and forms are discussed in section III.E.
Considerations related to important uncertainties inherent in such an
approach are discussed in section III.F. Advice from CASAC on such a
new standard is presented in section III.G. The Administrator's
proposed decisions on such a new standard are presented in section
III.H.

A. Ambient Air Indicators

In considering alternative ambient air indicators, the PA primarily
focuses on the important attribute of association. Association in a
broad sense refers to how well an ambient air indicator relates to the
ecological effects of interest by virtue of both the framework that
links the ambient indicator and effects and the empirical evidence that
quantifies the linkages. The PA also considers how measurable or
quantifiable an indicator is to enable its use as an effective
indicator of relevant ambient air concentrations.
As discussed above in section II.C, the PA concludes that
indicators other than NO2 and SO2 should be
considered as the appropriate indicators of oxides of nitrogen and
sulfur in the ambient air for protection against the acidification
effects associated with deposition of the associated nitrogen and
sulfur. This conclusion is based on the recognition that all forms of
nitrogen and sulfur in the ambient air contribute to deposition and
resulting acidification, and as such, NO2 and SO2
are incomplete indicators. In principle, the ambient indicators should
represent the species that are associated with oxides of nitrogen and
sulfur in the ambient air and can contribute acidifying deposition.
This includes both the species of oxides of nitrogen and sulfur that
are directly emitted as well as species transformed in the atmosphere
from oxides of nitrogen and sulfur that retain the nitrogen and sulfur
atoms from directly emitted oxides of nitrogen and sulfur. All of these
compounds are associated with oxides of nitrogen and sulfur in the
ambient air and can contribute to acidifying deposition.
The PA focuses in particular on the various compounds with nitrogen
or sulfur atoms that are associated with oxides of nitrogen and sulfur,
because the acidifying potential is specific to nitrogen and sulfur,
and not other atoms (e.g., H, C, O) whether derived from the original
source of oxides of nitrogen and sulfur emissions or from atmospheric
transformations. For example, the acidifying potential of each molecule
of NO2, NO, HNO3 or PAN is identical, as is the
potential for each molecule of SO2 or ion of particulate
sulfate, p-SO4. Each atom of sulfur affords twice the
acidifying potential of each atom of nitrogen.
1. Oxides of Sulfur
As discussed in the PA (US EPA, 2011, section 7.1.1), oxides of
sulfur include the gases sulfur monoxide (SO), SO2, sulfur
trioxide (SO3), disulfur monoxide (S2O), and
particulate-phase sulfur compounds (referred to as SO4) that
result from gas-phase sulfur oxides interacting with particles.
However, the sum of SO2 and SO4 does represent
virtually the entire ambient air mass of sulfur that contributes to
acidification. In addition to accounting for virtually all the
potential for acidification from oxidized sulfur in the ambient air,
there are reliable methods to monitor the concentrations of
SO2 and particulate SO4. In addition, much of the
data used to develop the technical basis for the standard developed in
the PA is based on monitoring or modeling of these species.\5\ The PA
concludes that the sum of SO2 and SO4, referred
to as SOX, are appropriate ambient air indicators of oxides
of sulfur because they represent virtually all of the acidification
potential of ambient air oxides of sulfur and there are reliable
methods suitable for measuring SO2 and SO4.
---------------------------------------------------------------------------

\5\ As discussed in chapter 2 of the PA, SO2 and
particulate SO4 are routinely measured in ambient air
monitoring networks, although only the Clean Air Status and Trends
Network (CASTNET) filter packs do not intentionally exclude particle
size fractions. The CMAQ treatment of SOX is the simple
addition of both species, which are treated explicitly in the model
formulation. All particle size fractions are included in the CMAQ
SOX estimates.
---------------------------------------------------------------------------

2. Oxides of Nitrogen
As discussed in the PA (US EPA, 2011, section 7.1.2),
NOy, as defined in chapter 2 of the PA, incorporates
basically all of the oxidized nitrogen species that have acidifying
potential and as such, NOy should be considered as an
appropriate indicator for oxides of nitrogen. Total reactive oxidized
nitrogen is an aggregate measure of NO and NO2 and all of
the reactive oxidized products of NO and NO2. That is,
NOy is a group of nitrogen compounds in which all of the
compounds are either an oxide of nitrogen or compounds in which the
nitrogen atoms came from oxides of nitrogen. Total reactive oxidized
nitrogen is especially relevant as an ambient indicator for
acidification in that it both relates to the oxides of nitrogen in the
ambient air and also represents the acidification potential of all
oxidized nitrogen species in the ambient air, whether an oxide of
nitrogen or derived from oxides of nitrogen.
There are currently available reliable methods of measuring
aggregate NOy. The term ``aggregate'' measure means that the
NOy, as measured, is not based on measuring each individual
species of NOy and calculating an NOy value by
summing the individual species. Rather, as described in chapter 2 of
the PA, current measurement techniques process all of the individual
NOy species to produce a single aggregate measure of all of
the nitrogen atoms associated with any NOy species.
Consequently, the NOy measurement effectively provides the
sum of all individual species, but the identity of the individual
species is lost. As discussed above, the accounting for the individual
nitrogen atoms is an accounting of the ambient air acidification
potential of oxides of nitrogen and their transformation products and
therefore the most relevant ambient indicator for aquatic acidification
effects associated with oxides of nitrogen.
This loss of the information on individual species motivated
consideration of alternative or more narrowly defined indicators for
oxides of nitrogen in the PA. Consideration of a subset of
NOy species was based on the following reasoning. First, the
actual dry deposition of nitrogen is determined on an individual
species basis by multiplying the species concentration times a species-
specific deposition velocity and then summed to develop an estimate of
total dry deposition. Consequently, the use of individual ambient
species has the potential to be more consistent with the underlying

[[Page 46113]]

science of deposition and, therefore, has the potential to allow for a
more rigorous evaluation of dry deposition with specialized field
studies. In addition, there has been a suggestion of focusing only on
the most quickly depositing NOy species, such as
HNO3, as contributions from other NOy species
such as NO2 may be negligible. These alternative indicators
are discussed below.
The PA considers the relative merits of using each individual
NOy species as part of a group of indicators. In so doing,
it was first noted that dry deposition of NOy is treated as
the sum of the deposition of each individual species in advanced
process-based air quality models like CMAQ, as described in chapter 2
of the PA. Conceptually one could extend this process-based approach by
using all NOy species individually as separate indicators
for oxides of nitrogen and requiring, for example, measurements of each
of the species, including the dominant species of HNO3,
particulate nitrate (p-NO3), true NO2, NO and
PAN. The potential attraction of using individual species would be the
reliance on actual deposition velocities. This could have more physical
meaning in comparison to a constructed model of aggregate deposition of
NOy, which is difficult to evaluate with observations
because of the assimilation of many species with disparate deposition
behavior. The PA notes that the major drawback of using individual
NOy species as the indicators is the lack of reliable
measurement techniques, especially for PAN and NO2 in rural
locations, which renders the use of virtually any individual
NOy species, except for NO and perhaps p-NO3, as
functionally inadequate from a measurement perspective.
The PA next considered the relative merits of using a subset of
NOy species as the indicators for oxides of nitrogen, as was
discussed above for oxides of sulfur. To the extent that certain
species provide relatively minor contributions to total NOy
deposition, it may be appropriate to consider excluding them as part of
the indicator. As discussed in chapter 2 of the PA, each nitrogen
species within the array of NOy species has species-specific
dry deposition velocities. For example, the deposition velocity of
HNO3 is much greater than the velocity for NO2
and, consequently, for a similar ambient air concentration,
HNO3 contributes more deposition of acidifying nitrogen
relative to NO2. In transitioning from source-oriented urban
locations to rural environments, the ratio of the concentrations of
HNO3 and PAN to NO2 increases.
Based on the reasoning that a larger fraction of the deposited
NOy is accounted for by total nitrate (the sum of
HNO3 and p-NO3), a surrogate for the more rapidly
depositing fraction of NOy, combined with the availability
of reliable total nitrate measurements through the CASTNET, the PA
considered using total nitrate as the indicator for oxides of nitrogen
(US EPA, 2011, appendix E). Nitrate would be expected to correlate well
with total reactive oxidized nitrogen deposition relative to
NOy (US EPA, 2011, chapter 2) despite the inherent noise
associated with variable contributions of low deposition velocity
species (e.g., NO2) that may have relatively high ambient
concentrations. However, modeling simulations suggest that
NOy may be a more robust indicator, relative to
HNO3, in terms of relating absolute changes in ambient air
concentrations to changes in nitrogen deposition driven by changes in
ambient concentrations of oxides of nitrogen (US EPA, 2011, Figure 2-
32).
Based on the above considerations, the PA concludes that
NOy should be considered as the appropriate ambient
indicator for oxides of nitrogen based on its direct relationship to
oxides of nitrogen in the ambient air and its direct relationship to
deposition associated with aquatic acidification. Because
NOy represents all of the potentially acidifying oxidized
nitrogen species in the ambient air, it is appropriately associated
with the deposition of potentially acidifying compounds associated with
oxides of nitrogen in the ambient air. In addition, there are reliable
methods available to measure NOy. Measurement of each
individual species of NOy, or the measurement of only a
subset of species of NOy, is less appropriate because there
are not reliable measurements methods available to measure all of the
individual species of NOy and a subset of species would fail
to account for significant portions of the oxidized reactive nitrogen
that relate to acidification.\6\
---------------------------------------------------------------------------

\6\ The PA also notes that NOy is a useful
measurement for model evaluation purposes, which is especially
important, recognizing the unique role that CMAQ plays in the
development of this standard, as described below in section III.B.
---------------------------------------------------------------------------

B. Form

Based on the evidence of the aquatic acidification effects caused
by the deposition of NOy and SOX, the PA (US EPA,
2011, section 7.2) presents the development of a new form that is
ecologically relevant for addressing such effects. The conceptual
design for the form of such a standard includes three main components:
an ecological indicator, deposition metrics that relate to the
ecological indicator, and a function that relates ambient air
indicators to deposition metrics. Collectively, these three components
link the ecological indicator to ambient air indicators, as illustrated
above in Fig II-1.
The simplified flow diagram in Figure II-1 compresses the various
atmospheric, biological, and geochemical processes associated with
acidifying deposition to aquatic ecosystems into a simplified
conceptual picture. The ecological indicator (left box) is related to
atmospheric deposition through biogeochemical ecosystem models (middle
box), which associate a target deposition load to a target ecological
indicator. Once a target deposition is established, associated
allowable air concentrations are determined (right box) through the
relationships between concentration and deposition that are embodied in
air quality models such as CMAQ. The following discussion describes the
development and rationale for each of these components, as well as the
integration of these components into the full expression of the form of
the standard using the concept of a national AAI that represents a
target ANC level as a function of ambient air concentrations. Spatial
aggregation issues associated with defining each of the terms of this
index are also addressed below.
The AAI is designed to be an ecologically relevant form of the
standard that determines the levels of NOy and
SOX in the ambient air that would achieve a target ANC limit
for the U.S. The intent of the AAI is to weight atmospheric
concentrations of oxides of nitrogen and sulfur by their propensity to
contribute to acidification through deposition, given the fundamental
acidifying potential of each pollutant, and to take into account the
ecological factors that govern acid sensitivity in different
ecosystems. The index also accounts for the contribution of reduced
nitrogen to acidification. Thus, the AAI encompasses those attributes
of specific relevance to protecting ecosystems from the acidifying
potential of ambient air concentrations of NOy and
SOX.
1. Ecological Indicator
In considering alternative ecological indicators, the PA again
primarily focuses on the attribute of association. In the case of an
ecological indicator for aquatic acidification, association refers to
the relationship between the indicator and adverse effects as discussed
in section II. Because of the conceptual structure of the form of an

[[Page 46114]]

AAI-based standard (Figure III-1), this particular ecological indicator
must also link up in a meaningful and quantifiable manner with
acidifying atmospheric deposition. In effect, the ecological indicator
for aquatic acidification is the bridge between biological impairment
and deposition of NOy and SOX.
This section presents the rationale in the PA for selecting ANC as
the appropriate ecological indicator for consideration. Recognizing
that ANC is not itself the causative or toxic agent for adverse aquatic
acidification effects, the rationale for using ANC as the relevant
ecological indicator is based on the following:
(1) The ANC is directly associated with the causative agents, pH
and dissolved Al, both through empirical evidence and mechanistic
relationships;
(2) Empirical evidence shows very clear and strong relationships
between adverse effects and ANC;
(3) The ANC is a more reliable indicator from a modeling
perspective, allowing use of a body of studies and technical analyses
related to ANC and acidification to inform the development of the
standard; and
(4) The ANC literally embodies the concept of acidification as
posed by the basic principles of acid base chemistry and the
measurement method used to estimate ANC and, therefore, serves as a
direct index to protect against acidification.
Ecological indicators of acidification in aquatic ecosystems can be
chemical or biological components of the ecosystem that are altered by
the acidifying effects of nitrogen and sulfur deposition. A desirable
ecological indicator for aquatic acidification is one that is
measurable or estimable, linked causally to deposition of nitrogen and
sulfur, and linked causally, either directly or indirectly to
ecological effects known or anticipated to adversely affect public
welfare.
As summarized in chapter 2 of the PA, atmospheric deposition of
NOy and SOX causes aquatic acidification through
the input of strong acid anions (e.g., NO3^- and
SO4^2-) that ultimately shifts the water chemistry
equilibrium toward increased hydrogen ion levels (or decreased pH). The
anions are deposited either directly to the aquatic ecosystem or
indirectly via transformation through soil nitrification processes and
subsequent drainage from terrestrial ecosystems. In other words, when
these anions are mobilized in the terrestrial soil, they can leach into
adjacent water bodies. Aquatic acidification is indicated by changes in
the surface water chemistry of ecosystems. In turn, the alteration of
surface water chemistry has been linked to negative effects on the
biotic integrity of freshwater ecosystems. There is a suite of chemical
indicators that could be used to assess the effects of acidifying
deposition on lake or stream acid-base chemistry. These indicators
include ANC; alkalinity (ALK); base neutralizing capacity, commonly
referred to as acidity (ACY); surface water pH; concentrations of
trivalent aluminum, Al\+3\; and concentrations of major anions
(SO4^2-, NO3^-), cations
(Ca\2+\, Mg\+2\, K\+\), or sums of cations or anions.
The ANC and ALK are very similar quantities and are used
interchangeably in the literature and for some of the analyses
presented in this document. Both ANC and ALK are defined as the amount
of strong acid required to reach a specified equivalence point. For
acid-base solutions, an equivalence point can be thought of as the
point at which the addition of strong acids (i.e., titration) is no
longer neutralized by the solution. This explains the term acid
neutralizing capacity, or ANC, as ANC relates directly to the capacity
of a system to neutralize acids. The differences between ANC and ALK
are based on operational definitions and subject to various
interpretations. The ANC is preferred over ALK as the body of
scientific evidence has focused on ANC and effects relationships. The
ALK is more widely associated with more general characterizations of
water quality such as the relative hardness of water associated with
carbonates.
Indictors such as the concentrations of specific anions, cations,
or their groupings, while relevant to acidification processes, are not
robust acidification indicators as it is the relative balance of
cations and anions that is more directly associated with acidification.
That balance is captured by ANC and ALK. Acidity, ACY, is the converse
of ANC and indicates how much strong base it takes to reach an
equivalence point. Because ACY is not used in most ecosystem
assessments, the body of information relating ACY to effects is too
limited to serve as a basis for an appropriate ecological indicator.
Aluminum and other metals are causative toxic agents that directly
impair biological functions. However, Al, or metals in general, have
high variability in concentrations that can be linked to effects, often
at extremely low levels which in some cases approach detectability
limits, exhibit rapid transient responses, and are often confounded by
the presence of other toxic metals. These concerns limit the use of
metals as reliable and measurable ecological indicators. Hydrogen ion
(H\+\) concentrations, using their negative logarithmic values, or pH,
are well correlated with adverse effects, as discussed above in section
II.A, and determine the solubility of metals such as aluminum. However,
pH is not a preferred acidification indicator due to its highly
transient nature and other concerns, as discussed below.
Having reasoned that ANC is a preferred indicator to ALK, ACY,
individual metals or groupings of ions, the PA considers the relative
merits of ANC compared to pH, which is a well recognized indicator of
acidity and a more direct causative agent with regard to adverse
effects. First, the linkage between ANC and pH is considered in
recognition of the causative association between pH and effects.
The ANC is not the direct causative toxic agent impacting aquatic
species diversity. The scientific literature generally emphasizes the
links between pH and adverse effects as described above in section
II.A. It is important, therefore, to consider the extent to which ANC
and pH are well related from a mechanistic perspective as well as
through empirical evidence. The ANC and pH are co-dependent on each
other based on the requirement that all solutions are electrically
neutral, meaning that any solution must satisfy the condition that all
negatively charged species must be balanced by all positively charged
species. The ANC is defined as the difference between strong anions and
cations (US EPA, 2011, equation 7-13).
While the chemistry can be complex, the co-dependency between ANC
and pH is explained by recognizing that positively charged hydrogen,
H\+\, is incorporated in the charge balance relationships related to
the overall solution chemistry which also defines ANC. The positive,
directional co-dependency (i.e., ANC and pH increase together) is
further explained in concept as ANC reflects how much strong acid
(i.e., how much hydrogen ion) it takes to titrate to an equivalence
point. Strong observed correlations between pH and ANC as described in
the PA support these mechanistic relationships.
As discussed above in section II.A, there are well established
examples of ANC correlating strongly with a variety of ecological
effects which are summarized in the PA (US EPA, 2011, Table 3-1).
Because pH and ANC are well correlated and linearly dependent over the
pH ranges (4.5-6) where adverse ecological effects are observed,
evidence of clear associations exist between ANC and adverse ecological
effects as described in the PA. In large measure, this dependence
between pH

[[Page 46115]]

and ANC and the relationship of both pH and ANC to effects, speak
directly to the appropriateness of ANC with respect to its use as an
ecological indicator.
Thus, there is a clear association between ANC and ecological
effects, although there is a more direct causal relationship between pH
and ecological effects. Nonetheless, ANC is preferred as an ecological
indicator based on its superior ability to provide a linkage with
deposition in a meaningful and quantifiable manner, a role that is
served far more effectively by ANC than by pH. While both ANC and pH
are clearly associated with the effects of concern, ANC is superior in
linking these effects to deposition.
The PA notes that the basis for this conclusion is that acidifying
atmospheric deposition of nitrogen and sulfur is a direct input of
potential acidity (ACY), or, in terms of ANC, such deposition is
relevant to the major anions that reduce the capacity of a water body
to neutralize acidity. Consequently, there is a well defined linear
relationship between potential acidifying deposition and ANC. This ANC-
deposition relationship facilitates the linkage between ecosystem
models that calculate an ecological indicator and the atmospheric
deposition of NOy and SOX. On the other hand,
there is no direct linear relationship between deposition and pH. While
acid inputs from deposition lower pH, the relationship can be extremely
nonlinear and there is no direct connection from a modeling or mass
balance perspective between the amount of deposition entering a system
and pH. The term ``mass balance'' underlies the basic formulation of
any physical modeling construct, for atmospheric or aquatic systems,
and refers to the accounting of the flow of mass into a system, the
transformation to other forms, and the loss due to flow out of a system
and other removal processes. The ANC is a conserved property. This
means that ANC in a water body can be accounted for by knowledge of how
much ANC initially exists, how much flows in and is deposited, and how
much flows out. In contrast, hydrogen ion concentration in the water,
the basis for pH, is not a conserved property as its concentration is
affected by several factors such as temperature, atmospheric pressure,
mixing conditions of a water body, and the levels of several other
chemical species in the system. The disadvantage of pH lacking
conservative properties is that there is a very complex connection
between changes in ambient air concentrations of NOy and
SOX and pH.
The discussion of basic water chemistry of natural systems in
chapter 2 of the PA provides further details on why pH is not a
conserved quantity and is subject to rapid transient response behavior
that makes it difficult to use as a reliable and functional ecological
indicator. The observed pH-to-ANC relationship (US EPA, 2011, figure 7-
2) partially explains the concern with pH responding too abruptly. In
the region where pH ranges roughly from 4.5 to 6 and is of greatest
relevance to effects (US EPA, 2011, figure 7-4), there clearly is more
sensitivity of pH to changes in ANC in the ANC range from approximately
0 to 50 [micro]eq/L. A focus on this part of the ANC-to-pH relationship
shows that ANC associates well with pH in a fairly linear manner.
However, the pH range from 4.5 to 6 also includes one of the very
steepest parts of the slope relating pH as a function of ANC, where ANC
ranges down below 0 [micro]eq/L, which is subject to very rapid change
in ANC, or deposition inputs. This part of the relationship coincides
with reduced levels of ANC and hence with reduced ability to neutralize
acids and moderate pH fluctuations. This response behavior can be
extended to considering how pH would change in response to deposition,
or ambient concentrations, of NOy and SOX, which
can be viewed as ``ANC-like'' inputs.
In summary, because ANC clearly links both to biological effects of
aquatic acidification as well as to acidifying inputs of NOy
and SOX deposition, the PA concludes that ANC is an
appropriate ecological indicator for relating adverse aquatic ecosystem
effects to acidifying atmospheric deposition of SOx and
NOy, and is preferred to other potential indicators. In
reaching this conclusion, the PA notes that in its review of the first
draft PA, CASAC concluded that ``information on levels of ANC
protective to fish and other aquatic biota has been well developed and
presents probably the lowest level of uncertainty in the entire
methodology'' (Russell and Samet, 2010a). In its more recent review of
the second draft PA, CASAC agreed ``that acid neutralizing capacity is
an appropriate ecological measure for reflecting the effects of aquatic
acidification'' (Russell and Samet, 2010b; p. 4).
2. Linking ANC to Deposition
There is evidence to support a quantified relationship between
deposition of nitrogen and sulfur and ANC. This relationship was
analyzed in the REA for two case study areas, the Adirondack and
Shenandoah Mountains, based on time-series modeling and observed
trends. In the REA analysis, long-term trends in surface water nitrate,
sulfate and ANC were modeled using MAGIC for the two case study areas.
These data were used to compare recent surface water conditions in 2006
with preindustrial conditions (i.e. preacidification 1860). The results
showed a marked increase in the number of acid impacted lakes,
characterized as a decrease in ANC levels, since the onset of
anthropogenic nitrogen and sulfur deposition, as discussed in chapter 2
of the PA.
In the REA, more recent trends in ANC, over the period from 1990 to
2006, were assessed using monitoring data collected at the two case
study areas. In both case study areas, nitrate and sulfate deposition
decreased over this time period. In the Adirondack Mountains, this
corresponded to a decreased concentration of nitrate and sulfate in the
surface waters and an increase in ANC (U.S. EPA, 2009, section
4.2.4.2). In the Shenandoah Mountains, there was a slight decrease in
nitrate and sulfate concentration in surface waters corresponding to
modest increase in ANC from 50 [mu]eq/L in 1990 to 67 [mu]eq/L in 2006
(U.S. EPA, 2009, section 4.2.4.3, Appendix 4, and section 3.4).
In the REA, the quantified relationship between deposition and ANC
was investigated using ecosystem acidification models, also referred to
as acid balance models or critical loads models (U.S. EPA, 2011,
section 2 and U.S. EPA, 2009, section 4 and Appendix 4). These models
quantify the relationship between deposition of nitrogen and sulfur and
the resulting ANC in surface waters based on an ecosystem's inherent
generation of ANC and ability to neutralize nitrogen deposition through
biological and physical processes. A critical load is defined as the
amount of acidifying atmospheric deposition of nitrogen and sulfur
beyond which a target ANC is not reached. Relatively high critical load
values imply that an ecosystem can accommodate greater deposition
levels than lower critical loads for a specific target ANC level.
Ecosystem models that calculate critical loads form the basis for
linking deposition to ANC.
As discussed in chapter 2 of the PA, both dynamic and steady state
models calculate ANC as a function of ecosystem attributes and
atmospheric nitrogen and sulfur deposition, and can be used to
calculate critical loads. Steady state models are time invariant and
reflect the long term consequences associated with an ecosystem
reaching equilibrium under a constant level of atmospheric deposition.
Dynamic models are time variant and take into account the time
dependencies inherent in ecosystem hydrology, soil and

[[Page 46116]]

biological processes. Dynamic models like MAGIC can provide the time
series response of ANC to deposition whereas steady state models
provide a single ANC relationship to any fixed deposition level.
Dynamic models naturally are more complex than steady state models as
they attempt to capture as much of the fundamental biogeochemical
processes as practicable, whereas steady state models depend on far
greater parameterization and generalization of processes that is
afforded, somewhat, by not having to accounting for temporal
variability.
The PA notes that steady state models are capable of addressing the
question of what does it take to reach and sustain a specific level of
ANC. Dynamic models are also capable of addressing that question, but
can also address the question of how long it takes to achieve that
result. Dynamic models afford the ability for more comprehensive
treatment of a variety of processes throughout the surface, soil and
bedrock layers within an ecosystem. For example, steady state models
treat sulfate as a mobile anion throughout the system, meaning that the
sulfate that is deposited to a watershed enters the water column and is
not influenced by soil adsorption or cation exchange. Dynamic models
can incorporate these time variant processes. The use of a steady state
model treating sulfate as totally mobile does not necessarily conflict
with the possibility of sulfate acting as a less than mobile ion at
certain times. The steady state assumption is premised on the long term
behavior of sulfate which can undergo periods of net adsorption
followed by periods of net desorption which can balance out over time.
The PA recognizes that as the richness of the available data increases,
in terms of parameters and spatial resolution, the incorporation of
dynamic modeling approaches in the standard setting process should
become more feasible. In determining an appropriate modeling approach
for the development of a NAAQS in this review, the PA considers both
the relevance of the question addressed as well as the ability to
perform modeling that provides relevant information for geographic
areas across the country.
Dynamic models require a large amount of catchment level-specific
data relative to steady state models. Because of the time invariant
nature of steady state models, the data requirements that integrate
across a broad spectrum of ecosystem processes is achievable and
available now at the national level. Water quality data to support
steady state models currently exist for developing a national data base
for modeling nearly 10,000 catchments in the contiguous U.S. In
contrast, the data needs to support dynamic models for national-scale
analyses simply are not available at this time. Further, the
information provided by steady state modeling would be sufficient to
develop and analyze alternative NAAQS and the kind of protection they
would afford. While it would be of interest to also obtain information
about how much time it would take for a target ANC level to be
achieved, the absence of such information does not preclude developing
and evaluating alternative NAAQS using the AAI structure. Based on the
above considerations, the PA concludes that at this time steady state
critical load modeling is an appropriate tool for linking long-term ANC
levels to atmospheric deposition of nitrogen and sulfur for development
of an AAI that has national applicability.
A steady state model is used to define the critical load, which is
the amount of atmospheric deposition of nitrogen (N) and sulfur (S)
beyond which a target ANC is not achieved and sustained.\7\ It is
expressed as:
---------------------------------------------------------------------------

\7\ This section discusses the linkages between deposition of
nitrogen and sulfur and ANC. Section III.B.3 then discusses the
linkages between atmospheric concentrations of NOY and
SOX and deposition of nitrogen and sulfur.

[GRAPHIC] [TIFF OMITTED] TP01AU11.024

---------------------------------------------------------------------------
Where:

CLANClim(N + S) is the critical load of deposition, with
units of equivalent charge/(area-time);
[BC]0^-* is the natural contribution of
base cations from weathering, soil processes and preindustrial
deposition, with units of equivalent charge/volume;
[ANClim] is the target ANC value, with units of
equivalent charge/volume; Q is the catchment level runoff rate
governed by water mass balance and dominated by precipitation, with
units of distance/time; and
Neco is the amount of nitrogen deposition that is effectively
neutralized by a variety of biological (e.g., nutrient uptake) and
physical processes, with units of equivalent charge/(area-time).

Equation III-1 is a modified expression that adopts the basic
formulation of the steady state models that are described in chapter 2
of the PA. More detailed discussion of the rationale, assumptions and
derivation of equation III-1, as well as all of the equations in this
section, are included in Appendix B of the PA. The equation simply
reflects the amount of deposition of nitrogen and sulfur from the
atmosphere, CLANClim(N + S), that is associated with a
sustainable long-term ANC target, [ANClim], given the
capacity of the natural system to generate ANC,
[BC]0^-*, and the capacity of the natural system
to neutralize nitrogen deposition, Neco. This expression of critical
load is valid when nitrogen deposition is greater than Neco.\8\ The
runoff rate, Q, allows for balancing mass in the two environmental
mediums--atmosphere and catchment. This critical load expression can be
focused on a single water system or more broadly. To extend
applicability of the critical load expression (equation III-1) from the
catchment level to broader spatial areas, the terms Qr and
CLr, are used, which are the runoff rate and critical load,
respectively, of the region over which all the atmospheric terms in the
equation are defined.
---------------------------------------------------------------------------

\8\ Because Neco is only relevant to nitrogen deposition, in
rare cases where Neco is greater than the total nitrogen deposition,
the critical load would be defined only in terms of acidifying
deposition of sulfur and the Neco term in equation III-1 would be
set to zero.
---------------------------------------------------------------------------

In considering the contributions of SOx or
NOy species to acidification, it is useful to think of every
depositing nitrogen atom as supplying one equivalent charge unit and
every sulfur atom as depositing two charge units. The PA uses
equivalent charge per volume as a normalizing tool in place of the more
familiar metrics such as mass or moles per volume. This allows for a
clearer explanation of many of the relationships between atmospheric
and ecosystem processes that incorporate mass and volume unit
conventions somewhat specific to the environmental media of concern
(e.g., m\3\ for air and liter for liquid water). Equivalent charge
reflects the chemistry equilibrium fundamentals that assume
electroneutrality, or balancing charge where the sum of cations always
equals the sum of anions.
As presented above, the terms S and N in the CLANClim (N
+ S) term broadly represent all species of sulfur or nitrogen that can
contribute to

[[Page 46117]]

acidifying deposition. This follows conventions used in the scientific
literature that addresses critical loads, and it reflects all possible
acidifying contributions from any sulfur or nitrogen species. For all
practical purposes, S reflects SOx as described above, the
sum of sulfur dioxide gas and particulate sulfate. However, N in
equation III-1 includes both oxidized forms, consistent with the
ambient indicator, NOy, in addition to the reduced nitrogen
species, ammonia and ammonium ion, referred to as NHx. The
NHX is included in the critical load formulation because it
contributes to potentially acidifying nitrogen deposition.
Consequently, from a mass balance or modeling perspective, the form of
the standard needs to account for NHX, as described below.
3. Linking Deposition to Ambient Air Indicators
The last major component of the form illustrated in Figure III-1
addresses the linkage between deposition of nitrogen and sulfur and
concentrations of the ambient air indicators, NOY and
SOX. To link ambient air concentrations with deposition, the
PA defines a transference ratio, T, as the ratio of total wet and dry
deposition to ambient concentration, consistent with the area and time
period over which the standard is defined. To express deposition of
NOY and SOX in terms of NOY and
SOX ambient concentrations, two transference ratios were
defined, where TSOx equals the ratio of the combined dry and
wet deposition of SOx to the ambient air concentration of
SOx, and TNOY equals the ratio of the combined
dry and wet deposition of NOY to the ambient air
concentration of NOY.
As described in chapter 7 of the PA, reduced forms of nitrogen
(NHx) are included in total nitrogen in the critical load
equation, III-1. Reduced forms of nitrogen are treated separately, as
are NOy and SOx, and the transference ratios are
applied. This results in the following critical load expression that is
defined explicitly in terms of the indicators NOY and
SOx:
[GRAPHIC] [TIFF OMITTED] TP01AU11.025

This is the same equation as III-1, with the deposition associated with
the critical load translated to deposition from ambient air
concentrations via transference ratios. In addition, deposition of
reduced nitrogen, oxidized nitrogen and oxidized sulfur are treated
separately.
Transference ratios are a modeled construct, and therefore cannot
be compared directly to measurable quantities. There is an analogy to
deposition velocity, as a transference ratio is basically an aggregated
weighted average of the deposition velocities of all contributing
species across dry and wet deposition, and transference ratio units are
expressed as distance/time. However, wet deposition commonly is not
interpreted as the product of a concentration times a velocity. Direct
wet deposition observations are available which integrate all of the
processes, regardless of how well they may be understood, related to
wet deposition into a measurable quantity. There are reasonable
analogies between the processes governing dry and wet deposition, from
a fundamental mass transfer perspective. In both cases there is a
transfer of mass between the dry ambient phase and another medium,
either a surface or vegetation in the case of dry deposition, or a rain
droplet or cloud in the case of wet precipitation. The specific
thermodynamic properties and chemical/biological reactions that govern
the transfer of dry mass to plants or aqueous droplets differ, but
either process can be based on conceptualizing the product of a
concentration, or concentration difference, times a mass transfer
coefficient which is analogous to the basic dry deposition model: dry
deposition = concentration x velocity (U.S. EPA, 2011, Appendix F).
Transference ratios require estimates of wet deposition of
NOy and SOX, dry deposition of NOY and
SOX, and ambient air concentrations of NOY and
SOx. Possible sources of information include model estimates
or a combination of model estimates and observations, recognizing that
dry deposition is a modeled quantity that can use observed or modeled
estimates of concentration. The limited amount of NOY
measurements in acid-sensitive areas as well as the combination of
representative NOY, SO2 and SO4
observations generally preclude the use of observations for development
of a standard that is applicable nationally.
The PA considers a blending of observations and models to take
advantage of their relative strengths; e.g., combining the NADP wet
deposition observations, modeled dry deposition, and a mix of modeled
and observed concentrations, using the model for those species not
measured or measured with very sparse spatial coverage. A potential
disadvantage of mixing and matching observations and model estimates is
to lose consistency afforded by using just modeling alone. A modeling
platform like CMAQ is based on adhering to consistent treatment of mass
conservation, by linking emission inputs with air concentrations and
concentrations to deposition. Inconsistencies from combining processes
from different analytical platforms increase the chance that mass (of
nitrogen or sulfur) would unintentionally be increased or decreased as
the internal checking that assures mass conservation is lost.
Transference ratios incorporate a broad suite of atmospheric processes
and consequently an analytical approach that instills consistency in
the linkage of these processes is preferable to an approach lacking
such inherent consistency. This contention does not mean that
observations alone, if available, could not be used, but suggests that
the inconsistencies in combining models and observations for the
purposes of developing transference ratios has the potential for
creating unintended artifacts.
While there is a reasonable conceptual basis for the concept of an
aggregated deposition velocity referred to in the PA as a transference
ratio, there is very limited ability to compare observed and calculated
ratios. This is because the deposition velocity is dependent on
individual species, and the mass transfer processes of wet and dry
removal, while conceptually similar, are different. Consequently, there
does not exist a meaningful approach to measure such an aggregated or
lumped parameter. Therefore, at this time, the evaluation of
transference ratios is based on sensitivity studies, analysis of
variability, and comparisons with other models, as described in
Appendix F of the PA.
As discussed in Appendix F, the interannual variability, as well as
the sensitivity to emission changes of roughly 50 percent, results in
changes of transference ratios of approximately 5 to 10 percent. Part
of the reason for this inherent stability is due to the co-dependence
of concentration and deposition. For example, as concentrations are
reduced as a result of emissions reductions, deposition in turn

[[Page 46118]]

is reduced since deposition is a direct linear function of
concentration leading to negligible impact on the deposition-to-
concentration ratio. Likewise, an overestimate of concentration likely
does not induce a bias in the transference ratio. While it is important
to continue to improve the model's ability to match ambient
concentrations in time and space, the bias of a modeled estimate of
concentration relative to observations does not necessarily result in a
bias in a calculated transference ratio. In effect, this consideration
of bias cancellation reduces the sensitivity of transference ratios to
model uncertainties and affords increased confidence in the stability
of these ratios. Based on the series of sensitivity and variability
analyses, the PA concludes that the transference ratios are relatively
stable and provide a sound metric for linking deposition and
concentration.
As discussed in the PA, transference ratios are dependent on the
platform upon which they are constructed. Comparisons of transference
ratios constructed from different modeling platforms do exhibit
significant differences. While this divergence of results may be
explained by a variety of differences in process treatments, input
fields and incommensurabilities in species definitions and spatial
configurations, it does suggest two very important conclusions. First,
the idea of using multiple platforms for different parts of the country
may be problematic as there does not exist a reliable approach to judge
acceptance which is almost always based on comparisons to observations.
Second, since transference ratios are based on concentrations and
deposition, as the uncertainties in each of those components are
reduced, the relative uncertainty in the ratios also is reduced. This
means that basic improvements in the model's ability to reproduce
observed wet deposition and ambient concentration fields enhance the
relative confidence in the constructed transference ratios. Similarly,
as in-situ dry deposition flux measurements become available that
enable a more rigorous evaluation and diagnosis of modeled dry
deposition processes, the expected improved treatment of dry deposition
also would increase confidence in transference ratios. Finally,
deposition is directly related to ambient air concentrations. Models
like CMAQ rely on the concentration-to-deposition linkage to calculate
deposition, which is the foundation for broadly based and robust
assessments addressing atmospheric deposition. In principle, the use of
a modeled constructed transference ratio is based on the same premise
by which we use models to estimate deposition in the first place.
The shortage of widely available ambient air observations and the
fact that estimates of dry deposition requires modeling, collectively
suggests that a unified modeling platform is the best approach for
constructing transference ratios. The PA (U.S. EPA, 2011, section 2)
considers CMAQ and other models, such as CAMx and Canada's AURAMS--A
Unified Regional Air-quality Modeling System (Smythe et al., 2008), and
concludes that CMAQ is the preferred modeling platform for constructing
transference ratios. This conclusion reflects the view that for the
purposes of defining transference ratios, a modeling platform should:
(1) Be a multiple pollutant model recognizing the myriad of connections
across pollutant categories that directly and indirectly impact
nitrogen and sulfur characterization, (2) include the most
comprehensive scientific treatments of atmospheric processes that
relate directly and indirectly to characterizing concentrations and
deposition, (3) have an infrastructure capability that accommodates the
inclusion of improved scientific treatments of relevant processes and
important input fields, and (4) undergo frequent reviews regarding the
adequacy of the underlying science as well as the appropriateness in
applications. The CMAQ platform exhibits all these characteristics. It
has been (and continues to be) extensively evaluated for several
pollutant categories, and is supported by a central infrastructure of
EPA scientists, whose mission is to improve and evaluate the CMAQ
platform. More directly, CMAQ, and its predecessor versions, has a long
track record going back to the NAPAP in the 1980s of specific
improvements in deposition processes, which are described in Appendix F
of the PA.
4. Aquatic Acidification Index
Having established the various expressions that link atmospheric
deposition of nitrogen and sulfur to ANC and the transference ratios
that translate atmospheric concentrations to deposition of nitrogen and
sulfur, the PA derived the following expression of these linkages,
which separates reduced forms of nitrogen, NHX, from
oxidized forms:
[GRAPHIC] [TIFF OMITTED] TP01AU11.026

Equation III-3 is the basic expression of the form of a standard
that translates the conceptual framework into an explicit expression
that defines ANC as a function of the ambient air indicators,
NOY and SOX reduced nitrogen deposition,\9\ and
the critical load necessary to achieve a target ANC level. This
equation calculates an expected ANC value based on ambient
concentrations of NOY and SOX. The calculated ANC
will differ from the target ANC (ANClim) depending on how much the
nitrogen and sulfur deposition associated with NOY,
SOX, and NHX differs from the critical load
associated with just achieving the target ANC.
---------------------------------------------------------------------------

\9\ Because NHx is characterized directly as deposition, not as
an ambient concentration in this equation, no transference ratio is
needed for this term.
---------------------------------------------------------------------------

Based on equation III-3, the PA defines an AAI that is more simply
stated using terms that highlight the ambient air indicators:
[GRAPHIC] [TIFF OMITTED] TP01AU11.027

where the AAI represents the long term (or steady state) ANC level
associated with ambient air concentrations of NOY and
SOX. The factors F1 through F4 convey three attributes: a
relative measure of the ecosystem's ability to neutralize acids (F1),
the acidifying potential of reduced nitrogen deposition (F2), and the
deposition-to-concentration translators for NOY (F3) and
SOX (F4).

Specifically:

F1 = ANClim + CLr/Qr;

[[Page 46119]]

F2 = NHx/Qr = NHx deposition divided by Qr;
F3 = TNOy/ Qr; TNOy is the transference ratio
that converts ambient air concentrations of NOy to deposition of
NOy; and
F4 = TSOx/ Qr; TSOx is the transference ratio
that converts ambient air concentrations of SOX to
deposition of SOX.

All of these factors include representative Qr to maintain unit (and
mass) consistency between the AAI and the terms on the right side of
equation III-4.
The F1 factor is the target ANC level plus the amount of deposition
(critical load) the ecosystem can receive and still achieve the target
level. It incorporates an ecosystem's ability to generate acid
neutralizing capacity through base cation supply ([BC]*0) and to
neutralize acidifying nitrogen deposition through Neco, both of which
are incorporated in the CL term. As noted above, because Neco can only
neutralize nitrogen deposition (oxidized or reduced) there may be rare
cases where Neco exceeds the combination of reduced and oxidized
nitrogen deposition. Consequently, to ensure that the AAI equation is
applicable in all cases that may occur, equation III-4 is conditional
on total nitrogen deposition, {NHX + F3[NOy]{time} , being
greater than Neco. In rare cases where Neco is greater than
{NHX + F3[NOy]{time} , F2, F3, and Neco would be set equal
to 0 in the AAI equation. The consequence of setting F2 and F3 to zero
is simply to constrain the AAI calculation just to SOx, as
nitrogen would have no bearing on acidifying contributions in this
case.
The PA concludes that equation III-4 (U.S. EPA, 2011,equation 7-
12), which defines the AAI, is ecologically relevant and appropriate
for use as the form of a national standard designed to provide
protection for aquatic ecosystems from the effects of acidifying
deposition associated with concentrations of oxides of nitrogen and
sulfur in the ambient air. This AAI equation does not, however, in
itself, define the spatial areas over which the terms of the equation
would apply. To specify values for factors F1 through F4, it is
necessary to define spatial areas over which these factors are
determined. Thus, it is necessary to identify an approach for spatially
aggregating water bodies into ecologically meaningful regions across
the U.S., as discussed below.
5. Spatial Aggregation
As discussed in the PA, one of the unique aspects of this form is
the need to consider the spatial areas over which values for the F
factors in the AAI equation are quantified. Ecosystems across the U.S.
exhibit a wide range of geological, hydrological and vegetation
characteristics that influence greatly the ecosystem parameters, Q,
BC0^-* and Neco that are incorporated in the AAI. Variations
in ecosystem attributes naturally lead to wide variability in the
sensitivities of water bodies in the U.S. to acidification, as well as
in the responsiveness of water bodies to changes in acidifying
deposition. Consequently, variations in ecosystem sensitivity, and the
uncertainties inherent in characterizing these variations, must be
taken into account in developing a national standard. In developing a
secondary NAAQS to protect public welfare, the focus of the PA is on
protecting sensitive populations of water bodies, not on each
individual water body, which is consistent with the Agency's approach
to protecting public health through primary NAAQS that focus on
susceptible populations, not on each individual.
The approach used for defining ecologically relevant regions across
the U.S. in the PA (U.S. EPA, 2011, section 7.2.5) is described below,
along with approaches to characterizing each region as acid sensitive
or relatively non-acid sensitive. This characterization facilitates a
more detailed analysis and focus on those regions that are relatively
more acid sensitive. This characterization is also used to avoid over-
protection in relatively non-acid sensitive regions, regions that would
receive limited benefit from reductions in the deposition of oxides of
nitrogen and sulfur with respect to aquatic acidification effects.
Approaches to developing representative values for each of the terms in
the AAI equation (factors F1 through F4) for each ecologically relevant
region for which sufficient data are available are then discussed.
These spatial aggregation approaches are generally applicable to the
contiguous U.S. The following discussion also addresses the development
of factors for data-limited regions and specifically for Hawaii, Alaska
and the U.S. territories.
Stated more simply, this section discusses appropriate ways to
divide the country into ecologically relevant regions; to characterize
each region as either acid sensitive or relatively non-acid sensitive;
and to determine values of factors F1 through F4 for each region,
taking into consideration the acid sensitivity of each region and the
availability of relevant data. For each such region, the AAI would be
calculated based on the values of factors F1 through F4 specified for
that region.
In considering approaches to spatial aggregation, the PA focuses on
methods that have been developed to define ecologically relevant
regions, referred to as ecoregions, which are meaningfully related to
the factors that are relevant to aquatic acidification. As noted above,
the PA did not focus on looking at each individual water body, nor did
it focus on aggregating over the entire nation, which would preclude
taking into account the inherent variability in atmospheric and
ecological factors that fundamentally modify the relationships that are
central to the development of an ecologically relevant AAI.
Based on considering available classification schemes, the PA
concludes that Omernik's ecoregion classification (as described at
http://www.epa.gov/wed/pages/ecoregions) is the most appropriate method
to consider for the purposes of this review. This classification offers
several levels of spatial delineation, has undergone an extensive
scientific peer review process, and has explicitly been applied to
delineating acid sensitive areas within the U.S. Further, the PA
concludes that ecoregion level III (Figure III-1) resolution, with 84
defined ecoregions in the contiguous U.S.,\10\ is the most appropriate
level to consider for this purpose. The spatial resolution afforded by
level III strikes an appropriate balance relative to the reasoning that
supports conclusions on indicators, as discussed above. The PA
concludes that the most detailed level of resolution (level IV) is not
appropriate given the limited data availability to address nearly 1,000
subdivisons within that level and the currently evolving nature of
level IV regions. Further, level III ecoregions are preferred to level
II in that level III ecoregions, but not level II ecoregions, are
largely contiguous in space which allows for a more coherent
development of information to quantify the AAI factors and to
characterize the concentrations of NOy and SOx in the ambient air
within each ecoregion.
---------------------------------------------------------------------------

\10\ We note that an 85th area within Omernik's Ecoregion Level
III is currently being developed for California.
---------------------------------------------------------------------------

Appendix C of the PA includes a description of each level III
ecoregion. The PA notes that the use of ecoregions is an appropriate
spatial aggregation scheme for an AAI-based standard focused on
deposition-related aquatic acidification effects, while many of the
same ecoregion attributes may be applicable in subsequent NAAQS reviews
that may address other deposition-related aquatic and terrestrial
ecological effects. Because atmospheric deposition is modified by
ecosystem attributes, the types of vegetation, soils, bedrock geology,
and

[[Page 46120]]

topographic features that are the basis of this ecoregion
classification approach also will likely be key attributes for other
deposition-related effects (e.g., terrestrial acidification, nutrient
enrichment) that link atmospheric concentrations to an aquatic or
terrestrial ecological indicator.
[GRAPHIC] [TIFF OMITTED] TP01AU11.028

a. Ecoregion Sensitivity
The PA used Omernik's original alkalinity data (U.S. EPA, 2011,
section 2) and more recent ANC data to delineate two broad groupings of
ecoregions: Acid-sensitive and relatively non-acid sensitive
ecoregions. This delineation was made to facilitate greater focus on
those ecoregions with water bodies that generally have greater acid
sensitivity and to avoid over-protection in regions with generally less
sensitive water bodies. The approach used to delineate acid-sensitive
and relatively non-acid sensitive regions included an initial
numerical-based sorting scheme using ANC data, which categorized
ecoregions with relatively high ANC values as being relatively non-acid
sensitive. This initial delineation resulted in 29 of the 84 Omernik
ecoregions being categorized as acid sensitive. Subsequently, land use
data were also considered to determine to what extent an ecoregion is
of a relatively pristine and rural nature by quantifying the degree to
which active management practices related to development and
agriculture occur in each ecoregion.
The overall objective is to produce a logical and practical
grouping of ecoregions that experience adverse conditions with respect
to aquatic acidification and are likely to respond to changes in
concentrations of NOy and SOx in the ambient air
and to the related deposition levels. To achieve this goal, a two-step
process has been applied, first identifying acid sensitive ecoregions
based on water quality data alone, and second identifying among those
acid-sensitive ecoregions those with highly developed and managed
areas. These highly developed and managed ecoregions are placed in a
non-acid sensitive category to avoid over protection beyond what is
requisite to protect public welfare. More specifically, in determining
an ecoregion's acid sensitivity status in step 1, ANC data across the
84 ecoregions are sorted (U.S. EPA, 2011, section 7) to determine the
number of water bodies within a region with ANC values suggestive of
acid sensitivity, so as to screen out regions with an overabundance of
high ANC values. In reviewing the ANC data, the PA identified 29
ecoregions that meet two criteria: (1) Greater than 5 percent of water
bodies with data with ANC values less than 200 [micro]eq/L and (2)
greater than 1 percent of water bodies with ANC values less than 100
[micro]eq/L. In step 2, land use data were used to identify those acid
sensitive ecoregions with significant managed areas that would not be
considered as having a relatively pristine and rural character. The
percentage of the combination of developed (residential,
transportation, industrial and commercial) and agricultural (croplands,
pastures, orchards, vineyards) land use was used as an indicator of
managed land use area. Forest cover was used as an indicator of non-
managed land use more directly reflecting the pristine quality of a
region. Based on the 2006 National Land Cover Data base (NLCD, http://www.epa.gov/mrlc/nlcd-2006.html), acid sensitive ecoregions would meet
both of the following land use data

[[Page 46121]]

criteria: Percent of developed and agricultural area less than 20
percent combined with forested area greater than 50 percent. The
combination of steps 1 and 2 identify 22 relatively acid sensitive
areas (Table III-1 and Figure III-2).
Consideration was also given to the use of naturally acidic
conditions in defining relatively non-acid sensitive areas. For
example, several of the ecoregions located in plains near the coast
exhibit elevated dissolved organic carbon (DOC) levels, which is
associated with naturally acidic conditions. The DOC in surface waters
is derived from a variety of weak organic acid compounds generated from
the natural availability and decomposition of organic matter from
biota. Consequently, high DOC is associated with ``natural'' acidity,
with the implication that a standard intended to protect against
atmospheric contributions to acidity is not an area of focus. The
evidence suggests that several of the more highly managed ecoregions in
coastal or near coastal transition zones are associated with relatively
high DOC values, typically exceeding on average 5 mg/l, compared to
other acid sensitive areas. Although there is sound logic to interpret
naturally acidic areas as relatively non-acid sensitive, natural
acidity indicators were not explicitly included in defining relatively
non-acid sensitive areas as there does not exist a consensus-based
quantifiable scientific definition of natural acidity. Approaches to
explicitly define natural acidity likely will be pursued in future
reviews of the standard.

Table III-1--List of 22 Acid-Sensitive Areas
------------------------------------------------------------------------
Ecoregion
Ecoregion name No.
------------------------------------------------------------------------
Ridge and Valley............................................ 8.4.1
Northern Appalachian Plateau and Uplands.................... 8.1.3
Piedmont.................................................... 8.3.4
Western Allegheny Plateau................................... 8.4.3
Southwestern Appalachians................................... 8.4.9
Boston Mountains............................................ 8.4.6
Blue Ridge.................................................. 8.4.4
Ouachita Mountains.......................................... 8.4.8
Central Appalachians........................................ 8.4.2
Northern Lakes and Forests.................................. 5.2.1
Maine/New Brunswick Plains and Hills........................ 8.1.8
North Central Appalachians.................................. 5.3.3
Northern Appalachian and Atlantic Maritime Highlands........ 5.3.1
Columbia Mountains/Northern Rockies......................... 6.2.3
Middle Rockies.............................................. 6.2.10
Wasatch and Uinta Mountains................................. 6.2.13
North Cascades.............................................. 6.2.5
Cascades.................................................... 6.2.7
Southern Rockies............................................ 6.2.14
Sierra Nevada............................................... 6.2.12
Idaho Batholith............................................. 6.2.15
Canadian Rockies............................................ 6.2.4
------------------------------------------------------------------------

[GRAPHIC] [TIFF OMITTED] TP01AU11.029

[[Page 46122]]

b. Representative Ecoregion-Specific Factors
Having concluded that the Omernik level III ecoregions are an
appropriate approach to spatial aggregation for the purpose of a
standard to address deposition-related aquatic acidification effects,
the PA uses those ecoregions to define each of the factors in the AAI
equation. As discussed below, factors F1 through F4 in equation III-4
are defined for each ecoregion by specifying ecoregion-specific values
for each factor based on monitored or modeled data that are
representative of each ecoregion.
i. Factor F1
As discussed above, factor F1 reflects a relative measure of an
ecosystem's ability to neutralize acidifying deposition, and is defined
as: F1 = ANClim + CLr/Qr. The value of F1 for each ecoregion would be
based on a representative critical load for the ecoregion
(CLr) associated with a single national target ANC level
(ANClim, discussed below in section III.D), as well as on a
representative runoff rate (Qr). To specify ecoregion-
specific values for the term Qr, the PA used the median
value of the distribution of Q values that are available for water
bodies within each ecoregion. To specify ecoregion-specific
representative values for the term CLr in factor F1, a
distribution \11\ of calculated critical loads was created for the
water bodies in each ecoregion for which sufficient water quality and
hydrology data are available.\12\ The representative critical load was
then defined to be a specific percentile of the distribution of
critical loads in the ecoregion. Thus, for example, using the 90th
percentile means that within an ecoregion, 90 percent of the water
bodies would be expected to have higher calculated critical loads than
the representative critical load. That is, if the representative
critical load were to occur across the ecoregion, 90 percent of the
water bodies would be expected to achieve the national ANC target or
better.
---------------------------------------------------------------------------

\11\ The distribution of critical loads was based on CL values
calculated with Neco at the lake level. Consideration could also be
given to using a distribution of CLs without Neco and adding the
ecoregion average Neco value to the nth percentile critical load.
This would avoid cases where the lake-level Neco value potentially
could be greater than total nitrogen deposition. The CL at the lake
level represents the CL for the lake to achieve the specified
national target ANC value.
\12\ The PA judged the data to be sufficient for this purpose if
data are available from more than 10 water bodies in an ecoregion.
---------------------------------------------------------------------------

The specific percentile selected as part of the definition of F1 is
an important parameter that directly impacts the representative
critical load specified for each ecoregion, and therefore the degree of
protectiveness of the standard. A higher percentile corresponds to a
lower critical load and, therefore, to lower allowable ambient air
concentrations of NOy and SOx and related
deposition to achieve a target AAI level. In conjunction with the other
terms in the AAI equation, alternative forms can be appropriately
characterized in part by identifying a range of alternative
percentiles. The choice of an appropriate range of percentiles to
consider for acid-sensitive and relatively non-acid sensitive
ecoregions, respectively, is discussed below.
a. Acid-Sensitive Ecoregions
In identifying percentiles that are appropriate to consider for the
purpose of calculating factor F1 for ecoregions characterized as acid
sensitive, the PA concludes that it is appropriate to focus on the
lower (more sensitive) part of the distribution of critical loads, so
as to ensure that the ecoregion would be represented by relatively more
acid sensitive water bodies within the ecoregion. Specifying factor F1
in this way would help to define a standard that would be protective of
the population of acid sensitive water bodies within an ecoregion,
recognizing that even ecoregions characterized as acid sensitive may
contain a number of individual water bodies that are not acid
sensitive. The PA recognizes that there is no basis for independently
evaluating the degree of protectiveness afforded by any specific
percentile value, since it is the combination of form and level, in
conjunction with the indicator and averaging time, which determine the
degree of protectiveness of a standard. In light of this, the PA
concludes that it is appropriate to consider a range of percentiles,
from well above the 50th percentile, or median, of the distribution to
somewhat below the highest value (in terms of sensitivity; a high
degree of sensitivity corresponds to a low value for critical load).
More specifically, the PA concludes it is appropriate to consider
percentiles in the range of the 70th to the 90th percentile (of
sensitivity). This conclusion is based on the judgment that it would
not be appropriate to represent an ecoregion with the lowest or near
lowest critical load, so as to avoid potential extreme outliers that
can be seen to exist at the extreme end of the data distributions,
which would not be representative of the population of acid sensitive
water bodies within the ecoregion and could lead to an overly
protective standard. At the same time, in considering ecoregions that
are inherently acid sensitive, it is judged to be appropriate to limit
the lower end of the range for consideration to the 70th percentile, a
value well above the median of the distribution, so that a substantial
majority of acid-sensitive water bodies are protected.
In considering this conclusion, the CASAC Panel noted that the data
bases for calculating critical loads within an ecoregion are not
necessarily representative of all water bodies within an ecoregion.
That is, in many ecoregions the lake sampling design used in studies
that generated the relevant data may have focused on the relatively
more sensitive water bodies within an ecoregion (Russell and Samet,
2011a). Consequently, a given percentile of the distribution of
calculated critical loads, based on sampled water bodies, may not be
representative of that percentile of all water bodies across an entire
ecoregion. To the extent that the sampling of water bodies within an
ecoregion was skewed toward the relatively more sensitive water bodies,
selecting a given percentile from the distribution of available
critical loads would result in a somewhat higher percentile of all
water bodies within that ecoregion having a higher calculated critical
load than the representative critical load value. Thus, the extent to
which study sampling designs have resulted in skewed distributions of
calculated critical loads is an uncertainty that is appropriate to
consider in selecting a percentile for the purpose of defining the
factor F1 in the AAI equation.
b. Non-Acid Sensitive Ecoregions
With regard to identifying percentiles that are appropriate to
consider for the purpose of calculating factor F1 for ecoregions
characterized as relatively non-acid sensitive, the PA recognizes that
while such ecoregions are generally less sensitive to acidifying
deposition from oxides of nitrogen and sulfur, they may contain a
number of water bodies that are acid sensitive. This category includes
ecoregions that are well protected from acidification effects due to
natural production of base cations and high ANC levels, as well as
naturally acidic systems with limited base cation production and
consequently very low critical loads. Therefore, the use of a critical
load that would be associated with a highly sensitive water body in a
naturally acidic system would impose a high degree of relative
protection in terms of allowable ambient air concentrations of oxides
of nitrogen and sulfur and

[[Page 46123]]

related deposition, while potentially affording little or no public
welfare benefit from attempting to improve a naturally acidic system.
Based on these considerations, the PA concludes it is appropriate
to consider the use of a range of percentiles that extends lower than
the range identified above for acid-sensitive ecoregions. Consideration
of a lower percentile would avoid representing a relatively non-acid
sensitive ecoregion by a critical load associated with relatively more
acid-sensitive water bodies. In particular, the PA concludes it is
appropriate to focus on the median or 50th percentile of the
distribution of critical loads so as to avoid over-protection in such
ecoregions. Recognizing that relatively non-acid sensitive ecoregions
generally are not sampled to the extent that acid-sensitive ecoregions
are, it also is appropriate to consider using the median critical load
of all relatively non-acid sensitive ecoregions for each such
ecoregion.
ii. Factor F2
As discussed above, factor F2 is the amount of reduced nitrogen
deposition within an ecoregion, including the deposition of both
ammonia gas and ammonium ion, and is defined as: F2 = NHX/
Qr. The PA calculated the representative runoff rate, Qr,
using a similar approach as noted above for factor F1; i.e., the median
value of the distribution of Q values that are available for water
bodies within each ecoregion. In the PA, 2005 CMAQ model simulations
over 12-km grids are used to calculate an average value of
NHX for each ecoregion. The NHX term is based on
annual average model outputs for each grid cell, which are spatially
averaged across all the grid cells contained in each ecoregion to
calculate a representative annual average value for each ecoregion. The
PA concludes that this approach of using spatially averaged values is
appropriate for modeling, largely due to the relatively rapid mixing of
air masses that typically results in relatively homogeneous air quality
patterns for regionally dispersed pollutants. In addition, there is
greater confidence in using spatially averaged modeled atmospheric
fields than in using modeled point-specific fields.
This averaging approach is also used for the air concentration and
deposition terms in factors F3 and F4, as discussed below. The PA notes
that modeled NHX deposition exhibits greater spatial
variability than the other modeled terms in factors F3 and F4.
Recognizing this greater variability, the PA concludes that it would be
appropriate to consider alternative approaches to specifying the value
of NHX. One such approach might involve the use of more
localized and/or contemporaneous modeling in areas where this term is
likely to be particularly variable and important.
iii. Factors F3 and F4
As discussed above, factors F3 and F4 are the ratios that relate
ambient air concentrations of NOy and SOX to the
associated deposition, and are defined as follow: F3 = TNOy/
Qr and F4 = TSOx/ Qr. TNOy is the transference
ratio that converts ambient air concentrations of NOy to
deposition of NOy and TSOx is the transference
ratio that converts ambient air concentrations of SOX to
deposition of SOX. The representative runoff rate,
Qr, is calculated as for factors F1 and F2. The transference
ratios are based on the 2005 CMAQ simulations, using average values for
each ecoregion, as noted above for factor F2. More specifically, the
transference ratios are calculated as the annual deposition of
NOy or SOX spatially averaged across the
ecoregion and divided by the annual ambient air concentration of
NOy or SOX, respectively, spatially averaged
across the ecoregion.
c. Factors in Data-Limited Ecoregions
As discussed above in section III.B.5.a, in the PA the initial
delineation of acid-sensitive and relatively non-acid sensitive
ecoregions was based on available ANC and alkalinity data. Areas not
meeting the ANC criteria described above are categorized as relatively
non-acid sensitive. The development of a reasonable distribution of
critical loads for water bodies within an ecoregion for the purpose of
identifying the representative critical load requires additional data,
including more specific water quality data for major cations and
anions. This means that the water bodies that can be used to develop a
distribution of critical loads is generally a subset of those water
bodies for which ANC data are available Consequently, there are certain
ecoregions with sparse data that are not suitable for developing a
distribution of critical loads.
As noted above, the PA judges that it is not appropriate to develop
such distributions based on data from less than ten water bodies within
an ecoregion. Twelve such ecoregions, which included only relatively
non-acid sensitive ecoregions, were characterized as being data-
limited. For these ecoregions, the PA considered alternative approaches
to specifying values for the terms CLr and Qr for
the purpose of determining values for each of the factors in the AAI
equation. For these data-limited ecoregions, the PA judges that it is
appropriate to use the median values of CLr and
Qr from the distributions of these terms for all other
relatively non-acid sensitive ecoregions, rather than attempting to use
severely limited data to develop a value for these terms based solely
on data from such an ecoregion. Further, consideration could be given
to using a single national default value for all relatively non-acid
sensitive ecoregions. The PA notes that this data limitation is not a
concern in specifying values for the other terms in the AAI equation
for such ecoregions, since those terms are based on data from the 2005
CMAQ model simulation, which covers all ecoregions across the
contiguous U.S.
d. Application to Hawaii, Alaska, and the U.S. Territories
The above methods for specifying ecoregion-specific values for the
factors in the AAI equation apply to those ecoregions within the
contiguous U.S. For areas outside the continental U.S., including
Hawaii, Alaska, and the U.S. Territories, there is currently a lack of
available data to characterize the sensitivity of such areas, as well
as a lack of water body-specific data and CMAQ-type modeling to specify
values for the F1 through F4 factors. Thus, the PA has considered
possible alternative approaches to specifying values for factors F1
through F4 in the AAI equation for these areas.
One such approach could be to specify area-specific values for the
factors based on values derived for ecoregions with similar acid
sensitivities, to the extent that relevant information can be obtained
to determine such similarities. Such an approach would involve
conducting an analysis to characterize similarities in relevant
ecological attributes between ecoregions in the contiguous U.S. and
these areas outside the contiguous U.S. so as to determine the
appropriateness of utilizing ecoregion-specific values for the
CLr and Qr terms from one or more ecoregions
within the contiguous U.S. This approach would also involve conducting
additional air quality modeling for the areas that are outside the
geographical scope of the currently available CMAQ model simulations,
so as to develop the other information necessary to specify values for
factors F2 through F4 for these areas.
A second approach could rely on future data collection efforts to
establish relevant ecological data within these areas that, together
with additional air quality modeling, could be used to specify area-
specific values for factors

[[Page 46124]]

F1 through F4. Until such time as relevant data become available, these
areas could be treated the same as data-limited ecoregions in the
contiguous U.S. that are relatively non-acid sensitive.
The PA concludes that either approach would introduce substantial
uncertainties that arise from attempting to extrapolate values based on
similarity assumptions or arbitrarily assigning values for factors in
the AAI equation that would be applicable to these areas outside the
contiguous U.S. In light of such uncertainties, the PA concludes that
it would also be appropriate to consider relying on the existing
NO2 and SO2 secondary standards in these areas
for protection of any potential direct or deposition-related ecological
effects that may be associated with the presence of oxides of nitrogen
and sulfur in the ambient air. The PA concludes that relying on
existing secondary standards in these areas is preferable to using a
highly uncertain approach to allow for the application of a new
standard based on the AAI in the absence of relevant area-specific
data.
6. Summary of the AAI Form
With regard to the form of a multi-pollutant air quality standard
to address deposition-related aquatic acidification effects, the PA
concludes that consideration should be given to an ecologically
relevant form that characterizes the relationships between the ambient
air indicators for oxides of nitrogen and sulfur, the related
deposition of nitrogen and sulfur, and the associated aquatic
acidification effects in terms of a relevant ecological indicator.
Based on the available information and assessments, consideration
should be given to using ANC as the most appropriate ecological
indicator for this purpose, in that it provides the most stable metric
that is highly associated with the water quality properties that are
directly responsible for the principal adverse effects associated with
aquatic acidification: Fish mortality and reduced aquatic species
diversity.
The PA developed such a form, using a simple equation to calculate
an AAI value in terms of the ambient air indicators of oxides and
nitrogen and sulfur and the relevant ecological and atmospheric factors
that modify the relationships between the ambient air indicators and
ANC. Recognizing the spatial variability of such factors across the
U.S., the PA concludes it is appropriate to divide the country into
ecologically relevant regions, characterized as acid-sensitive or
relatively non-acid-sensitive, and specify the value of each of the
factors in the AAI equation for each such region. Omernik ecoregions,
level III, are identified as the appropriate set of regions over which
to define the AAI. There are 84 such ecoregions that cover the
continental U.S. This set of ecoregions is based on grouping a variety
of vegetation, geological, and hydrological attributes that are
directly relevant to aquatic acidification assessments and that allow
for a practical application of an aquatic acidification standard on a
national scale.
The PA defines AAI by the following equation: AAI = F1 - F2 -
F3[NOy] - F4[SOX]. Factors F1 through F4 would be
defined for each ecoregion by specifying ecoregion-specific values for
each factor based on monitored or modeled data that are representative
of each ecoregion. The F1 factor is also defined by a target ANC value.
More specifically:
(1) F1 reflects a relative measure of an ecosystem's ability to
neutralize acidifying deposition. The value of F1 for each ecoregion
would be based on a representative critical load for the ecoregion
associated with a single national target ANC level, as well as on a
representative runoff rate. The representative runoff rate, which is
also used in specifying values for the other factors, would be the
median value of the distributions of runoff rates within the ecoregion.
The representative critical load would be derived from a distribution
of critical loads calculated for each water body in the ecoregion for
which sufficient water quality and hydrology data are available. The
representative critical load would be defined by selecting a specific
percentile of the distribution.
In identifying a range of percentiles that are appropriate to
consider for this purpose, regions categorized as acid sensitive were
considered separately from regions categorized as relatively non-acid
sensitive. For acid sensitive regions, the PA concludes that
consideration should be given to selecting a percentile from within the
range of the 70th to the 90th percentile. The lower end of this range
was selected to be appreciably above the median value so as to ensure
that the critical load would be representative of the population of
relatively more acid sensitive water bodies within the region, while
the upper end was selected to avoid the use of a critical load from the
extreme tail of the distribution which is subject to a high degree of
variability and potential outliers. For relatively non-acid sensitive
regions, the PA concludes that consideration should be given to
selecting the 50th percentile to best represent the distribution of
water bodies within such a region, or alternatively to using the median
critical load of all relatively non-acid sensitive areas, recognizing
that such areas are far less frequently evaluated than acid sensitive
areas. Using either of these approaches would avoid characterizing a
generally non-acid-sensitive region with a critical load that is
representative of relatively acid sensitive water bodies that may exist
within a generally non-acid sensitive region.
(2) F2 reflects the deposition of reduced nitrogen. Consideration
should be given to specifying the value of F2 for each region based on
the averaged modeled value across the region, using national CMAQ
modeling that has been conducted by EPA. Consideration could also be
given to alternative approaches to specifying this value, such as the
use of more localized and/or contemporaneous modeling in areas where
this term is likely to be particularly variable and important.
(3) F3 and F4 reflect transference ratios that convert ambient air
concentrations of NOy and SOX, respectively, into
related deposition of nitrogen and sulfur. Consideration should be
given to specifying the values for F3 and F4 for each region based on
CMAQ modeling results averaged across the region. We conclude that
specifying the values or the transference ratios based on CMAQ modeling
results alone is preferred to an alternative approach that combines
CMAQ model estimates with observational data.
(4) The terms [NOy] and [SOX] reflect ambient
air concentrations measured at monitoring sites within each region.
Using the equation, a value of AAI can be calculated for any
measured values of ambient NOy and SOX. For such
a NAAQS, the Administrator would set a single, national value for the
level of the AAI used to determine achievement of the NAAQS, as
discussed below in section III.D. The ecoregion-specific values for
factors F1 through F4 would be specified by EPA based on the most
recent data and CMAQ model simulations, and codified as part of such a
standard. These factors would be reviewed and updated as appropriate in
the context of each periodic review of the NAAQS.
The PA developed specific F factors for each ecoregion based on the
approach discussed above, using alternative percentiles and alternative
national target ANC levels. The results of this analysis for ecoregions
characterized as acid sensitive are presented in Table 7-1a-d in the
PA.

[[Page 46125]]

C. Averaging Time

As discussed in section 7.3 of the PA, aquatic acidification can
occur over both long- and short-term timescales. Long-term cumulative
deposition of nitrogen and sulfur is reflected in the chronic acid-base
balance of surface waters as indicated by measured annual ANC levels.
Similarly, the use of steady state critical load modeling, which
generates critical loads in terms of annual cumulative deposition of
nitrogen and sulfur, means that the focus of ecological effects studies
based on critical loads is on the long-term equilibrium status of water
quality in aquatic ecosystems. Much of the evidence of adverse
ecological effects associated with aquatic acidification, as discussed
above in section II.A, is associated with chronically low ANC levels.
Protection against a chronic ANC level that is too low is provided by
reducing overall annual average deposition levels for nitrogen and
sulfur.
Reflecting this focus on long-term acidifying deposition, the PA
developed the AAI that links ambient air indicators to deposition-
related ecological effects, in terms of several factors, F1 through F4.
As discussed above, these factors are all calculated as annual average
values, whether based on water quality and hydrology data or on CMAQ
model simulations. In the context of a standard defined in terms of the
AAI, the PA concludes that it is appropriate to consider the same
annual averaging time for the ambient air indicators as is used for the
factors in the AAI equation.
We also recognize that short-term (i.e., hours or days) episodic
changes in water chemistry, often due to changes in the hydrologic flow
paths, can have important biological effects in aquatic ecosystems.
Such short-term changes in water chemistry are termed ``episodic
acidification.'' Some streams may have chronic or base flow chemistry
that is generally healthy for aquatic biota, but may be subject to
occasional acidic episodes with potentially lethal consequences. Thus,
short-term episodic ecological effects can occur even in the absence of
long-term chronic acidification effects.
Episodic declines in pH and ANC are nearly ubiquitous in drainage
waters throughout the eastern U.S. Episodic acidification can result
from several mechanisms related to changes in hydrologic flow paths.
For example, snow can store nitrogen deposited throughout the winter
and snowmelt can then release this stored nitrogen, together with
nitrogen derived from nitrification in the soil itself, in a pulse that
leads to episodic acidification in the absence of increased deposition
during the actual episodic acidification event. The PA notes that
inputs of nitrogen and sulfur from snowpack and atmospheric deposition
largely cycle through soil. As a result, short-term direct deposition
inputs are not necessarily important in episodic acidification. Thus,
as noted in chapter 3 of the ISA, protection against episodic acidity
events can be achieved by establishing a higher chronic ANC level.
Taken together, the above considerations support the conclusion
that it is appropriate to consider the use of a long-term average for
the ambient air indicators NOy and SOX for an
aquatic acidification standard defined in terms of the AAI. The use of
an annual averaging time for NOy and SOX
concentrations would be appropriate to provide protection against low
chronic ANC levels, which in turn would protect against both long-term
acidification and acute acidic episodes.
The PA has also considered interannual variability in both ambient
air quality and in precipitation, which is directly related to the
deposition of oxides of nitrogen and sulfur from the ambient air. While
ambient air concentrations show year-to-year variability, often the
year-to-year variability in precipitation is considerably greater,
given the highly stochastic nature of precipitation. The use of
multiple years over which annual averages are determined would dampen
the effects of interannual variability in both air quality and
precipitation. For the ambient air indicators, the use of multiple-year
averages would also add stability to calculations used to judge whether
an area meets a standard defined in terms of the AAI. Consequently, the
PA concludes that an annual averaging time based on the average of each
year over a consecutive 3- to 5-year period is appropriate to consider
for the ambient air indicators NOy and SOX. In
reaching this conclusion, the PA notes that in its comments on the
second draft PA, CASAC agreed that a 3- to 5-year averaging time was
appropriate to consider (Russell and Samet, 2010b).

D. Level

As discussed above, the PA concludes that ANC is the ecological
indicator best suited to reflect the sensitivity of aquatic ecosystems
to acidifying deposition from oxides of nitrogen and sulfur in the
ambient air. The ANC is an indicator of the aquatic acidification
expected to occur given the natural buffering capacity of an ecosystem
and the loadings of nitrogen and sulfur resulting from atmospheric
deposition. Thus, the PA developed a new standard for aquatic
acidification that is based on the use of chronic ANC as the ecological
indicator as a component in the AAI.
The level of the standard would be defined in terms of a single,
national value of the AAI. The standard would be met at a monitoring
site when the multi-year average of the calculated annual values of the
AAI was equal to or above the specified level of the standard.\13\ The
annual values of the AAI would be calculated based on the AAI equation
using the assigned ecoregion-specific values for factors F1 through F4
and monitored annual average NOy and SOX
concentrations. Since the AAI equation is based on chronic ANC as the
ecological indicator, the level chosen for the standard would reflect a
target chronic ANC value. As noted above, the assigned F factors for
each ecoregion would be determined by EPA in the rulemaking to set the
NAAQS, based on water quality and hydrology data, CMAQ modeling, the
selected percentile that is used to identify a representative critical
load within the ecoregion, and the selected level of the standard. The
combination of the form of the standard, discussed above in section
III.B, defined by the AAI equation and the assigned values of the F
factors in the equation, other elements of the standard including the
ambient air indicators (section III.A) and their averaging time
(section III.C), and the level of the standard determines the allowable
levels of NOy and SOX in the ambient air within
each ecoregion. All of the elements of the standard together determine
the degree of protection from adverse aquatic acidification effects
associated with oxides of nitrogen and sulfur in the ambient air. The
level of the standard plays a central role in determining the degree of
protection provided and is discussed below.
---------------------------------------------------------------------------

\13\ Unlike other NAAQS, where the standard is met when the
relevant value is at or below the level of the standard since a
lower standard level is more protective, in this case a higher
standard level is more protective.
---------------------------------------------------------------------------

The PA focuses primarily on information that relates degrees of
biological impairment associated with adverse ecological effects to
aquatic ecosystems to alternative levels of ANC in reaching conclusions
regarding the range of target ANC levels that is appropriate to
consider for the level of the standard. The PA develops the rationale
for identifying a range of target ANC levels that is appropriate to
consider by addressing questions related to the following areas: (1)
Associations between ANC and pH levels to provide an initial bounding
for the range of ANC

[[Page 46126]]

values to be considered; (2) evidence that allows for the delineation
of specific ANC ranges associated with varying degrees of severity of
biological impairment ecological effects; (3) the role of ANC in
affording protection against episodic acidity; (4) implications of the
time lag response of ANC to changes in deposition; (5) past and current
examples of target ANC values applied in environmental management
practices; and (6) data linking public welfare benefits and ANC.
1. Association Between pH Levels and Target ANC Levels
As discussed above in section II.A and more fully in chapter 3 of
the PA, specific levels of ANC are associated with differing levels of
risk of biological impairment in aquatic ecosystems, with higher levels
of ANC resulting in lower risk of ecosystem impacts, and lower ANC
levels resulting in risk of both higher intensity of impacts and a
broader set of impacts. While ANC is not the causal agent determining
biological effects in aquatic ecosystem, it is a useful metric for
determining the level at which a water body is protected against risks
of acidification. There is a direct correlation between ANC and pH
levels which, along with dissolved aluminum, are more closely linked to
the biological causes of ecosystem response to acidification.
Because there is a direct correlation between ANC and pH levels,
the selection of target ANC levels is informed in part through
information on effects of pH as well as direct studies of effects
related to ANC. Levels of pH are closely associated with ANC in the pH
range of approximately 4.5 to 7. Within this range, higher ANC levels
are associated with higher pH levels. At a pH level of approximately
4.5, further reductions in ANC generally do not correlate with pH, as
pH levels remain at approximately 4.5 while ANC values fall
substantially. Similarly, at a pH value of approximately 7, ANC values
can continue to increase with no corresponding increase in pH. As pH is
the primary causal indicator of effects related to aquatic
acidification, this suggests that ANC values below approximately -50
[mu]eq/L (the apparent point in the relationship between pH and ANC
where pH reaches a minimum) are not likely to result in further damage.
In addition, ANC values around and above approximately 100 [mu]eq/L
(the apparent region in the relationship where pH reaches a maximum)
are not likely to confer additional protection. As a result, the
initial focus in the PA was on target ANC values in the range of -50 to
100 [mu]eq/L.
2. ANC Levels Related to Effects on Aquatic Ecosystems
As discussed above in section II.A, the number of fish species
present in a water body has been shown to be positively correlated with
the ANC level in the water, with higher values supporting a greater
richness and diversity of fish species. The diversity and distribution
of phyto-zooplankton communities also are positively correlated with
ANC.
A summary of effects related to ANC ranges is shown above in Table
II-1. Within the ANC range from approximately -50 to 100 [mu]eq/L,
linear and sigmoidal relationships are observed between ANC and
ecosystem effects. On average, fish species richness is lower by one
fish species for every 21 [mu]eq/L decrease in ANC in Shenandoah
National Park streams (ISA, section 3.2.3.4). As shown in Table II-1,
ANC levels have been grouped into five categories related to expected
ecological effects, including categories of acute concern (<0 [mu]eq/
L), severe concern (0-20 [mu]eq/L), elevated concern (20-50 [mu]eq/L),
moderate concern (50-100 [mu]eq/L), and low concern (>100 [mu]eq/L).
This categorization is supported by a large body of research completed
throughout the eastern U.S. (Sullivan et al., 2006).
Water bodies with ANC values less than or equal to 0 [mu]eq/L at
based flow are chronically acidic. Such ANC levels can lead to complete
loss of species and major changes in the ability of water bodies to
support diverse biota, especially in water bodies that are highly
sensitive to episodic acidification. Based on the above considerations,
the PA has focused on target ANC levels no lower than 0 [mu]eq/L.
As discussed in the PA, biota generally are not harmed when ANC
values are >100 [mu]eq/L, due to the low probability that pH levels
will be below 7. In the Adirondacks, the number of fish species also
peaks at ANC values >100 [mu]eq/L. This suggests that at ANC levels
greater than 100 [mu]eq/L, little risk from acidification exists in
many aquatic ecosystems. At ANC levels below 100 [mu]eq/L, overall
health of aquatic communities can be maintained, although fish fitness
and community diversity begin to decline. At ANC levels ranging from
100 down to 50 [mu]eq/L, there is increasing likelihood that the
fitness of sensitive species (e.g., brook trout, zooplankton) will
begin to decline. When ANC concentrations are below 50 [mu]eq/L, the
probability of acidification increases substantially, and negative
effects on aquatic biota are observed, including large reductions in
diversity of fish species and changes in the health of fish
populations, affecting reproductive ability and fitness, especially in
water bodies that are affected by episodic acidification. While there
is evidence that ANC levels above 50 can confer additional protection
from adverse ecological effects associated with aquatic acidification
in some sensitive ecosystems, the expectation that such incremental
protection from adverse effects will continue up to an ANC level of 100
is substantially reduced. The PA concludes that the above
considerations support a focus on target ANC levels up to a level
greater than 50 [mu]eq/L but below 100 [mu]eq/L, such as up to a level
of 75 [mu]eq/L.
In considering the available scientific evidence, as summarized
here and discussed in more detail in the ISA and REA, in its review of
the second draft PA, CASAC expressed the following views about the
range of biological responses that corresponds to this range of ANC
levels (i.e., 0-100 [mu]eq/L):

There will likely be biological effects of acidification at
higher ANC values within this range, and there are relatively
insensitive organisms that are not impacted at ANC values at the low
end of this range. Adverse effects of acidification on aquatic biota
are fairly certain at the low end of this range of ANC and
incremental benefits of shifting waters to higher ANC become more
uncertain at higher ANC levels. There is substantial confidence that
there are adverse effects at ANC levels below 20 [mu]eq/L, and
reasonable confidence that there are adverse effects below 50
[mu]eq/L. Levels of 50 [mu]eq/L and higher would provide additional
protection, but the Panel has less confidence in the significance of
the incremental benefits as the level increases above 50 [mu]eq/L.
(Russell and Samet, 2010b)

The PA concludes that the above considerations, including the views
of CASAC, provide support for focusing on target ANC levels in the
range of 20 to 75 [mu]eq/L.
3. Consideration of Episodic Acidity
As discussed in the PA, across the broad range of ANC values from 0
to 100 [mu]eq/L, ANC affords protection against the likelihood of
decreased pH (and associated increases in Al) during long or short
periods. In general, the higher the ANC within this range, the lower
the probability of reaching low pH levels where direct effects such as
increased fish mortality occur, as shown in Table 3-1 of the PA.
Accordingly, greater protection would be achieved by target chronic ANC
values set high enough to avoid pH depression to levels associated with
elevated risk.

[[Page 46127]]

The specific relationship between ANC and the probability of
reaching pH levels of elevated risk varies by water body and fish
species. The ANC levels below 20 [mu]eq/L are generally associated with
high probability of low pH, leading to death or loss of fitness of
biota that are sensitive to acidification (US EPA, 2008, section
5.2.2.1; US EPA, 2009, section 5.2.1.2). At these levels, during
episodes of high acidifying deposition, brook trout populations may
experience lethal effects. In addition, the diversity and distribution
of zooplankton communities decline sharply at ANC levels below 20
[mu]eq/L. Overall, there is little uncertainty that significant effects
on aquatic biota are occurring at ANC levels below 20 [mu]eq/L.
It is clear that at ANC levels approaching 0 [mu]eq/L (Table II-1),
there is significant impairment of sensitive aquatic ecosystems with
almost complete loss of fish species. Avoiding ANC levels approaching 0
[mu]eq/L is particularly relevant to episodic spikes in acidity that
occur during periods of rapid snow melt and during and after major
precipitation events. Since the ANC range considered in the PA reflects
average, long-term base flow values, it is appropriate to consider
protecting against episodic drops in ANC values to a level as low as 0
[mu]eq/L. Staddard et al. (2003) noted on average a 30 [mu]eq/L
depression of ANC between spring and summer time values, indicating the
need to maintain higher base flow ANC levels to protect against ANC
levels below 0 [mu]eq/L. The above considerations do not provide
support for a target chronic ANC level as low as 0 [mu]eq/L for a
standard that would protect against significant harm to aquatic
ecosystems, including harm from episodic acidification. The PA
concludes that these considerations also support a lower end of the
range for consideration no lower than 20 [mu]eq/L.
The CASAC agreed with this conclusion in its comments on the second
draft PA (Russell and Samet, 2010b). The CASAC noted that ``there are
clear and marked biological effects at ANC values near 0 [mu]eq/L, so
this is probably not an appropriate target value'' for the AAI. With
regard to the likelihood of impairment of aquatic ecosystems due to
episodic acidification, in terms of specific target levels for chronic
ANC, CASAC expressed the following view:

Based on surface waters studied in the Northeast, decreases in
ANC associated with snowmelt [are] approximately 50 [mu]eq/L. Thus,
based on these studies, a long term ANC target level of 75 [mu]eq/L
would generally guard against effects from episodic acidification
down to a level of about 25 [mu]eq/L. (Russell and Samet, 2010b)

4. Consideration of Ecosystem Response Time
The PA notes that when considering a standard level to protect
against aquatic acidification, it is appropriate to take into account
both the time period to recovery as well as the potential for recovery
in acid-sensitive ecoregions. Ecosystems become adversely impacted by
acidifying deposition over long periods of time and have variable time
frames and abilities to recover from such perturbations. Modeling
presented in the REA (U.S. EPA, 2009, section 4.2.4) shows the
estimated ANC values for Adirondack lakes and Shenandoah streams under
pre-acidification conditions and indicates that for a small percentage
of lakes and streams, natural ANC levels would have been below 50
[mu]eq/L. Therefore, for these water bodies, reductions in acidifying
deposition are not likely to achieve an ANC of 50 [mu]eq/L or greater.
Conversely, for some lakes and streams the level of perturbation from
long periods of acidifying deposition has resulted in very low ANC
values compared to estimated natural conditions. For such water bodies,
the time to recovery would be largely dependent on future inputs of
acidifying deposition.
Setting a standard level in terms of a target chronic ANC level is
based on the long-term response of aquatic ecosystems. The time
required for a water body to achieve the target ANC level--given a
decrease in ambient air concentrations of NOy and
SOx and related acidifying deposition such that the critical
load for a target ANC is not exceeded--is often decades if not
centuries. In recognition of the potential public welfare benefits of
achieving the target ANC in a shorter time frame, the concept of target
loads had been developed. Target loads represent the depositional
loading that is expected to achieve a particular level of the
ecological indicator by a given time. For example, to achieve an ANC
level of 20 [mu]eq/L by 2030, it might be necessary to specify a higher
target ANC level of, for example, 50 [mu]eq/L, such that the
depositional loading would be reduced more quickly than would occur if
the depositional loading was based on achieving a target ANC level of
20 [mu]eq/L as a long-term equilibrium level. In this example, the
target ANC of 50 [mu]eq/L would ultimately be realized many years
later.
The above considerations have implications for selecting an
appropriate standard level, in that the standard level affects not only
the ultimate degree of protection that would be afforded by the
standard, but also the time frame in which such protection would be
realized. However, the PA recognizes that there is a great deal of
heterogeneity in response times among water bodies and that there is
only very limited information from dynamic modeling that would help to
quantify recovery time frames in areas across the country. As a
consequence, quantification of a general relationship between critical
loads associated with a specific long-term target ANC level and target
loads associated with achieving the target ANC level within a specific
time frame is not currently possible. Thus, while the time frame for
recovery is an important consideration in selecting an appropriate
range of levels to consider, the PA concludes that it can only be
considered in a qualitative sense at this time.
5. Prior Examples of Target ANC Levels
A number of regional organizations, states, and international
organizations have developed critical load frameworks to protect
against acidification of sensitive aquatic ecosystems. In considering
the appropriate range of target ANC levels for consideration in this
review, it is informative to evaluate the target ANC levels selected by
these different organizations, as well as the rationale provided in
support of the selected levels. Chapter 4 of the PA provides a detailed
discussion of how critical loads have been developed and used in other
contexts. Specific target values and their rationales are summarized
below.
The UNECE has developed critical loads in support of international
emissions reduction agreements. As noted in chapter 4 of the PA,
critical loads were established to protect 95 percent of surface waters
in Europe from an ANC less than 20 [micro]eq/L based on protection of
brown trout. Individual countries have set alternative ANC targets; for
example, Norway targets an ANC of 30 [micro]eq/L based on protection of
Atlantic salmon. Several states have established target ANC or pH
values related to protection of lakes and streams from acidification.
While recognizing that some lakes in the Adirondacks will have a
naturally low pH, the state of New York has established a target pH
value of 6.5 for lakes that are not naturally below 6.5. As noted
above, this level is associated with an ANC value that is likely to be
between 20 and 50 [micro]eq/L or possibly higher. New Hampshire and
Vermont have set ANC targets of 60 [micro]eq/L and 50 [micro]eq/L,
respectively. Tennessee has established site-specific target ANC

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values based on assessments of natural acidity, with a default value of
50 [micro]eq/L when specific data are not available.
Taken together, these policy responses to concerns about ecological
effects associated with aquatic acidification indicate that target ANC
values between 20 and 60 [mu]eq/L have been selected by states and
other nations to provide protection of lakes and streams in some of the
more sensitive aquatic ecosystems.
6. Consideration of Public Welfare Benefits
The point at which effects on public welfare become adverse is not
defined in the CAA. Characterizing a known or anticipated adverse
effect to public welfare is an important component of developing any
secondary NAAQS. According to the CAA, welfare effects include:

* * * effects on soils, water, crops, vegetation, manmade materials,
animals, wildlife, weather, visibility, and climate, damage to and
deterioration of property, and hazards to transportation, as well as
effect on economic values and on personal comfort and well-being,
whether caused by transformation, conversion, or combination with
other air pollutants. (CAA, section 302(h)).

Consideration of adversity to public welfare in the context of the
secondary NAAQS for oxides of nitrogen and sulfur can be informed by
information about losses in ecosystem services associated with
acidifying deposition and the potential economic value of those losses,
as summarized above in section II.C and discussed more fully in chapter
4 of the PA.
Ecosystem service losses at alternative ANC levels are difficult to
enumerate. However, in general there are categories of ecosystem
services, discussed in chapter 4 of the PA, that are related to the
specific ecosystem damages expected to occur at alternative ANC levels.
Losses in fish populations due to very low ANC (below 20 [mu]eq/L) are
likely associated with significant losses in value for recreational and
subsistence fishers. Many acid sensitive lakes are located in areas
with high levels of recreational fishing activity. For example, in the
northeastern U.S., where nearly 8 percent of lakes are considered
acidic, more than 9 percent of adults participate in freshwater
fishing, with an estimated value of approximately $5 billion in 2006.
This suggests that improvements in lake fish populations may be
associated with significant recreational fishing value.
As discussed in the PA, inland surface waters also provide cultural
services such as aesthetic and existence value and educational
services. To the extent that piscivorous birds and other wildlife are
harmed by the absence of fish in these waters, hunting and birdwatching
activities are likely to be adversely affected. A case study of the
value to New York residents of improving the health of lakes in the
Adirondacks found significant willingness to pay for those
improvements. When scaled to evaluate the improvement in lake health
from achieving ANC values of either 20 or 50 [mu]eq/L, the study
implies benefits to the New York population roughly on the order of
$300-900 million per year (in constant 2007$). The survey administered
in this study recognized that participants were thinking about the full
range of services provided by the lakes in question--not just the
recreational fishing services. Therefore the estimates of willingness
to pay include resident's benefits for potential hunting and
birdwatching activities and other ancillary services. These results are
just for New York populations. The PA concludes that if similar
benefits exist for improvements in other acid sensitive lakes, the
economic value to U.S. populations could be very substantial,
suggesting that, at least by one measure of impact on public welfare,
impacts associated with ANC less than 50 [mu]eq/L may be adverse to
public welfare.
7. Summary of Alternative Levels
Based on all the above considerations, the PA concludes that
consideration should be given to a range of standard levels from 20 to
75 [mu]eq/L. The available evidence indicates that target ANC levels
below 20 [mu]eq/L would be inadequate to protect against substantial
ecological effects and potential catastrophic loss of ecosystem
function in some sensitive aquatic ecosystems. While ecological effects
occur at ANC levels below 50 [mu]eq/L in some sensitive ecosystems, the
degree and nature of those effects are less significant than at levels
below 20 [mu]eq/L. Levels at and above 50 [mu]eq/L would be expected to
provide additional protection, although uncertainties regarding the
potential for additional protection from adverse ecological effects are
much larger for target ANC levels above about 75 [mu]eq/L, as effects
are generally appreciably less sensitive to changes in ANC at such
higher levels.
In reaching this conclusion in the PA, consideration was given to
the extent to which a target ANC level within this range would protect
against episodic as well as long-term ecological effects. Levels in the
mid- to upper-part of this range would be expected to provide greater
protection against short-term, episodic peaks in aquatic acidification,
while lower levels within this range would give more weight to
protection from long-term rather than episodic acidification.
Similarly, levels in the mid- to upper-part of this range would be
expected to result in shorter time periods for recovery given the lag
in ecosystem response in some sensitive ecosystems relative to levels
in the lower part of this range. The PA also notes that this range
encompasses target ANC values that have been established by various
States and regional and international organizations to protect against
acidification of aquatic ecosystems.
The PA recognizes that the level of the standard together with the
other elements of the standard, including the ambient air indicators,
averaging time, and form, determine the overall protectiveness of the
standard. Thus, consideration of a standard level should reflect the
strengths and limitations of the evidence and assessments as well as
the inherent uncertainties in the development of each of the elements
of the standard. The implications of considering alternative standards,
defined in terms of alternative combinations of levels and percentile
values that are a critical component of factor F1 in the form of the
standard, are discussed below in section III.E. Key uncertainties in
the various components of the standard are summarized and considered
below in section III.F.

E. Combined Alternative Levels and Forms

To provide some perspective on the implications of various
alternative multi-pollutant, AAI-based standards, the PA presented the
number of acid-sensitive ecoregions that would likely not meet various
sets of alternative standards. The alternative standards considered
were based on combinations of alternative target ANC levels, within the
range of 20 to 75 [mu]eq/L, and alternative forms, characterized by
alternative representative percentiles within the range of the 70th to
90th percentile. These alternative standards are also defined in terms
of the other elements of the standard: ambient air indicators NOy and
SOx, discussed above in section III.A; other elements of the form of
the standard, including ecoregion-specific values for factors F1
through F4 in the AAI equation, discussed above in section III.B.5; and
an annual averaging time for NOy and SOx,
discussed above in section III.C. With regard to the averaging time,
the assessment did not consider multi-year averaging of the calculated
annual AAI

[[Page 46129]]

values due to data limitations, including, for example, the lack of
CMAQ modeling for multiple consecutive years. In this assessment, we
characterize an ecoregion as likely not meeting a given alternative
standard if the calculated AAI value is less than the target ANC level
of the standard, recognizing that higher AAI values are more protective
than lower values.
The results of this assessment are presented in Table 7-1a-d in the
PA for a subset of ecoregions including those characterized as acid
sensitive. Calculated annual AAI values at the ecoregion level are
shown for each alternative standard considered. Based on these AAI
values, Table 7-2 in the PA summarizes the number of acid-sensitive
ecoregions that would likely not meet each of the alternative standards
considered.\14\ Calculated AAI values for all ecoregions categorized as
relatively non-acid sensitive are shown in Table D-5 in Appendix D of
the PA. In all cases, these relatively non-acid sensitive ecoregions
were estimated to meet all of the alternative standards considered in
this assessment.
---------------------------------------------------------------------------

\14\ Tables 7-1a-d and 7-2 in the PA present assessment results
for 29 ecoregions that had been initially characterized as acid
sensitive. Subsequently, based on a broader set of criteria used to
characterize ecoregions as acid sensitive, as discussed above in
section III.B.5.a, the set of ecoregions characterized as acid
sensitive was narrowed to include 22 ecoregions.
---------------------------------------------------------------------------

As described above, the AAI values presented in Table 7-1a-d of the
PA are based in part on data from 2005 CMAQ model simulations, which
was used to generate values for F2 through F4 in the AAI equation, as
well as to estimate annual average ambient air concentrations of
NOy and SOx that reflect recent air quality in
the absence of currently available monitored concentrations in
sensitive ecoregions across the country. Water quality and hydrology
data from water bodies within each ecoregion were also used in
calculating the AAI values. Such data were initially used to calculate
critical loads for each water body with sufficient data within an
ecoregion so as to identify the nth percentile critical load
representative of the ecoregion used in calculating the F1 factor for
the ecoregion. As expected, the number of ecoregions that likely would
not meet alternative standards increases with increasing percentile
values and target ANC levels (U.S. EPA, 2011, Table 7-2). Out of 22
acid-sensitive ecoregions, the number of ecoregions that would likely
not meet the alternative standards ranges from 22 for the most
protective alternative standard considered (75 [mu]eq/L, 90th
percentile) to 4 for the least protective alternative standard (20
[mu]eq/L, 70th percentile). It is apparent that both the percentile and
the level chosen have a strong influence, over the ranges considered,
in determining the number of areas that would likely not meet this set
of alternative standards.
The PA observes that there is one grouping of these acid-sensitive
ecoregions that would likely not meet almost all combinations of level
and form under consideration (U.S. EPA, 2011, Table 7-2 and Appendix
D). This group is made up of southern Appalachian mountain areas,
including North Central Appalachians, 5.3.3; Ridge and Valley, 8.4.1;
Central Appalachians, 8.4.2; Blue Ridge, 8.4.4; and Southwestern
Appalachians, 8.4.9. In addition, these ecoregions exhibit the highest
amounts of exceedance relative to alternative standards.
The Northern Appalachian and Atlantic Maritime Highlands (5.3.1),
which includes the Adirondacks, and the Northern Lakes and Forests
(5.2.1) of the upper midwest exhibit similar patterns with respect to
in the role of level and percentile in identifying regions not likely
to meet alternative standards, although there are considerably fewer
cases compared to the regions in the Appalachians.
In the mountainous west, the Sierra Nevada (6.2.12), Idaho
Batholith (6.2.15) and the Cascades (6.2.7) ecoregions likely would not
meet alternative standards in fewer cases relative to eastern regions,
with the Sierra Nevada ecoregion exhibiting relatively greater
sensitivity compared to all western regions. Only in the upper part of
the ranges of level and percentile do regions in the northern and
central Rockies likely not meet alternative standards.
In considering these findings, the PA observes that the standard as
defined by the AAI behaves in an intuitively logical manner. That is,
an increase in ecoregions likely not to meet the standard is associated
with higher alternative levels and percentiles, both of which
contribute to a lower regionally representative critical load.
Moreover, the areas of known adverse aquatic acidification effects are
identified, mostly in high elevation regions or in the northern
latitudes--the Adirondacks, Shenandoahs, northern midwest lakes and the
mountainous west. These results reflect the first application of a
nationwide model that integrates water quality and atmospheric
processes at a national scale and provides findings that are consistent
with our basic understanding of the extent of aquatic acidification
across the U.S. What is particularly noteworthy is that this model is
not initialized with a starting ANC based on water quality data, which
likely would result in a reproduction of water quality observations.
Rather, this standard reflects the potential of the changes in
atmospheric concentrations of NOy and SOx to
induce long-term sustained changes in surface water systems. The PA
notes that the fact that the patterns of adversity based on applying
this standard are commensurate with what is observed in surface water
systems provides confidence in the basic underlying formulation of the
standard.
The PA notes that the Appalachian mountain regions merit further
inspection as they stand out as areas with the largest relative
exceedances from a national perspective. Water quality data from these
regions as well as an emissions sensitivity CMAQ simulation were
considered to better understand the simulated behavior of these
regions. The maps and tables in appendix D of the PA include paired
comparisons of the CMAQ 2005 and emissions sensitivity simulations. The
emissions sensitivity simulation reflects domain-wide reductions in
NOy and SOx emissions of 48 percent and 42
percent, respectively, relative to 2005 base year emissions. The PA
assumes that this emissions sensitivity simulation is indicative of
future conditions.
The emissions sensitivity results project that many of the regions
that likely would not meet the alternative standards based on recent
air quality, especially at alternative levels of 20 and 35 [micro]eq/L,
would likely meet such standards in the future year scenario for the
Appalachian mountain regions. It is apparent that the AAI calculations
are especially sensitive to changes in SOx emissions as the
Appalachian regions have the highest SOx concentrations and
deposition rates (U.S. EPA, 2011,section 2), and the AAI equation
responds as expected to modeled reductions in SOx. The
emissions sensitivity scenario is a prospective application of the
standard, in the sense that rules derived from the air quality
management process result in reductions of NOy and
SOx emissions. Expected emission changes over the next two
decades should be far greater than the 42 percent and 48 percent,
respectively, SOx and NOy reductions used in this
analysis, with a consequent further reduction in areas that would
likely not meet alternative standards.
The Appalachian mountain regions generally have low DOC levels,
average runoff rates, moderately low base cation supply and highly
elevated sulfate concentrations. Collectively, those attributes do not
suggest naturally acidic conditions as the availability of

[[Page 46130]]

anthropogenic contributions of mineral acids is likely responsible for
observed low ANC values in those regions.
The PA notes the Sierra Nevada region as an interesting case study,
as it has some of the lowest critical load values nationally (U.S. EPA,
2011, Table D-3). Water quality data indicate extremely low sulfate, as
expected given the relatively low SO2 emissions in the
western U.S. Extremely low base cation supply and low Neco, which
mitigate the effect of nitrogen deposition, explain the low critical
load values. Low Neco values appear to associate well with high
elevation western U.S. regions, perhaps reflecting the more arid and
reduced vegetation density relative to eastern U.S. regions. The
proximity to high level nitrogen emissions combined with very low base
cation supply explains the cases where the Sierra region likely does
not meet alternative standards. Because Neco values are low in the
Sierras, the system responds effectively to reductions of
NOx emissions, as illustrated in the maps and tables of
Appendix D of the PA. Although Neco affords protection from the
acidifying effects of nitrogen deposition, the availability of
excessive nitrogen neutralization capacity also means that reductions
in nitrogen are not as effective as reductions in SOx in
reducing the calculated AAI.
In reviewing these results, the PA observes that the analysis of
the alternative combinations of level and form presented provide
context for considering the impact of different standards. Since the
AAI equation has been newly developed in the PA, these examples of
estimated exceedances help to address the question of whether the AAI
equation responds in a reasonable manner with regard to identifying
areas of concern and to prospective changes in atmospheric conditions
likely to result from future emissions reduction strategies. The PA
concludes that the behavior of the AAI calculations is both reasonable
and explainable, which the PA concludes serves to increase confidence
in considering a standard defined in terms of the AAI.

F. Characterization of Uncertainties

This section summarizes discussions of the results of analyses and
assessments, presented more fully in the PA (U.S. EPA, 2011, section
7.6 and Appendices F and G), intended to address the relative
confidence associated with the linked atmospheric-ecological effects
system described above. An overview of uncertainties is presented in
the context of the major structural components underlying the standard,
as well as with regard to areas of relatively high uncertainty. The
section closes with a discussion of data gaps and uncertainties
associated with the use of ecological and atmospheric modeling to
specify the factors in the AAI equation, which can be used to guide
future field programs and longer-term research efforts.
1. Overview of Uncertainty
As discussed in the PA (U.S. EPA, 2011, Table 7-3), there is
relatively low uncertainty with regard to the conceptual formulation of
the overall structure of the AAI-based standard that incorporates the
major associations linking biological effects to air concentrations.
Based on the strength of the evidence that links species richness and
mortality to water quality, the associations are strongly causal and
without any obvious confounding influence. The strong association
between the ecosystem indicator (ANC) and the causative water chemistry
species (dissolved aluminum and hydrogen ion) reinforces the confidence
in the linkage between deposition of nitrogen and sulfur and effects.
This strong association between ANC and effects is supported by a sound
mechanistic foundation between deposition and ANC. The same mechanistic
strength holds true for the relationship between ambient air levels of
nitrogen and sulfur and deposition, which completes the linkage from
ambient air indicators through deposition to ecological effects.
There are relatively higher uncertainties, however, in considering
specific elements within the structure of an AAI-based standard,
including the deposition of SOx, NOy, and
NHx as well as the critical load-related component, each of
which can vary within and across ecoregions. Overall system uncertainty
relates not just to the uncertainty in each such element, but also to
the combined uncertainties that result from linking these elements
together within the AAI-based structure. Some of these elements--
including, for example, dry deposition, pre-industrial base cation
production, and reduced nitrogen deposition--are estimated with less
confidence than other elements (U.S. EPA, 2011, Table 7.3). The
uncertainties associated with all of these elements, and the
combination of these elements through the AAI equation, are discussed
below and in the following sections related to measured data gaps and
modeled processes for both air quality and water quality.
The lack of observed dry deposition data is constrained by
resources and the lack of efficient measurement technologies. Progress
in reducing uncertainties in dry deposition will depend on improved
atmospheric concentration data and direct deposition flux measurements
of the relevant suite of NOy and SOx species.
Pre-industrial base cation productivity by definition is not
observable. Contemporary observations and inter-model comparisons are
useful tools that would help reduce the uncertainty in estimates of
preindustrial base cation productivity used in the AAI equation. In
characterizing contemporary base cation flux using basic water quality
measurements (i.e., major anion and cation species as defined in
equation 2.11 in the PA), it is reasonable to assume that a major
component of contemporary base cation flux is associated with pre-
industrial weathering rates. To the extent that multiple models
converge on similar solutions, greater confidence in estimating pre-
industrial base cation production would be achieved.
Characterization of NHx deposition has been evolving
over the last decade. The relatively high uncertainty in characterizing
NHx deposition is due to both the lack of field measurements
and the inherent complexity of characterizing NHx with
respect to source emissions and dry deposition. Because ammonia
emissions are generated through a combination of man-made and
biological activities, and ammonia is semi-volatile, the ability to
characterize spatial and temporal distributions of NHx
concentrations and deposition patterns is challenging. While direct
measurement of NHx deposition is resource intensive because
of the diffuse nature of sources (i.e., area-wide and non-point
sources), there have been more frequent deposition flux studies,
relative to other nitrogen species, that enable the estimation of both
emissions and dry deposition. Also, while ammonia has a relatively high
deposition velocity and traditionally was thought to deposit close to
the emissions release areas, the semi-volatile nature of ammonia
results in re-entrainment back into the lower boundary layer resulting
in a more dispersed concentration pattern exhibiting transport type
characteristics similar to longer lived atmospheric species. These
inherent complexities in source characterization and ambient
concentration patterns raise the uncertainty level of NHx in
general. However, the PA notes that progress is being made in measuring
ammonia with cost efficient samplers and anticipates the gradual
evolution of a spatially robust ammonia sampling network that would
help support analyses to reduce underlying uncertainties in
NHx

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deposition. Also, from an aquatic acidification perspective,
NHx is not as important a driver as NOy and
SOx in the mountainous areas in the eastern U.S. However,
the relative importance of NHx is likely to increase over
time, in light of air quality rules in place designed to reduce
emissions of NOy and SOx.
2. Uncertainties Associated With Data Gaps
In summarizing uncertainties with respect to available measurement
data and the use of ecological and atmospheric models, the PA
indentified data gaps and model uncertainties in relative terms by
comparing, for example, the relative richness of data between
geographic areas or environmental media. With regard to relevant air
quality measurements, the PA notes that such measurements are
relatively sparse in the western U.S. While the spatial extent of
CASTNET coverage has gradually incorporated western U.S. locations with
support from the NPS, the relative density of monitoring sites is much
less than that in the eastern U.S. This relative disparity in spatial
density of monitors is exacerbated as air quality patterns in the
mountainous west generally exhibit greater spatial heterogeneity due to
dramatic elevation gradients that impact meteorology and air mass flow
patterns. Similarly, water quality data coverage is far more
comprehensive in the eastern U.S. relative to the west
Measurements of NOy notably are lacking in both eastern
and western acid-sensitive ecoregions. This adds uncertainty to the use
of the AAI equation as the lack of NOy data limits efforts
to evaluate air quality modeling of NOy that is the basis
for quantifying factor F3 in the AAI equation. The lack of
NOy measurements also limits efforts to characterize the
variability and representativeness of modeled NOy
concentrations within and across ecoregions. Currently, the Agency's
ability to define the protection likely to be afforded by alternative
standards (in terms of alternative levels and percentiles) is
compromised by the lack of a full set of ambient air quality indicator
measurements, notably including NOy, throughout sensitive
ecoregions across the U.S.
Further, obtaining measurement of the dominant species that
comprise NOy (HNO3, true NO2, NO, p-
NO3, and PAN) would be useful to evaluate performance of
NOy samplers. Beyond the more well known dominant components
of NOy, research efforts would be needed to characterize
total reactive nitrogen that may include significant amounts of
organically-bound nitrogen (beyond PAN) which is poorly understood with
regard to emission sources and concentration levels.
Field measurements of NHx have been extremely limited,
but have begun to be enhanced through the NADP's passive ammonia
network (AMoN). The AMoN measures ammonia at over 50 sites, with more
than 35 at CASTNET locations. Enhanced spatial coverage of reduced
nitrogen measurements, particularly to understand within and across
ecoregion variability, and the inclusion of some continuous
observations would provide a better understanding of the uncertainty in
the F2 factor in the AAI equation and of the representativeness of
modeled NHx deposition within and across ecoregions.
With regard to water quality data, the PA notes that such data are
typically limited relative to air quality data sets, and are also
relatively sparse in the western U.S. The TIME/LTM water quality
sampling program in the eastern U.S. (as described in chapter 2 of the
PA) is an appropriate complement to national air monitoring programs as
it affords consistency across water bodies in terms of sampling
frequency and analysis protocols. Consideration should be given to
extending the TIME/LTM design to all acid sensitive ecoregions, with
priority for areas in the western mountains that are data limited and
showing initial signs of adversity particularly with respect to aquatic
acidification. The lack of a regulatory requirement for TIME/LTM often
jeopardizes funding support of this resource that is especially
valuable and cost effective. While there are several state and local
agency water quality data bases, it is unclear the extent to which
differences in sampling, chemical analysis and reporting protocols
would impact the use of such data for the purpose of better
understanding the degree of protectiveness that would be afforded by an
AAI-based standard within sensitive ecoregions across the country. In
addition, our understanding of water quality in Alaska and Hawaii and
the acid sensitivity of their ecoregions is particularly limited.
Water quality data and modeling support the standard setting
process. As more water bodies are sampled, the critical load data bases
would expand, enabling clearer delineation of ecoregion representative
critical loads in terms of the nth percentile. This would provide more
refined characterization of the degree of protection afforded by a
given standard. Longer term, the availability of water quality trend
data (annual to monthly sampled) would support accountability
assessments that examine if an ecoregion's response to air management
efforts is as predicted by earlier model forecasting. The most obvious
example is the long-term response of water quality ANC change to
changes in calculated AAI, deposition, ambient NOy and
SOx concentrations, and emissions. In addition, water
quality trends data provide a basis for evaluating and improving the
parameterizations of processes in critical load models applied at the
ecoregion scale related to nitrogen retention and base cation supply. A
better understanding of soil processes, especially in the southern
Appalachians, would enhance efforts to examine the variability within
ecoregions of the soil-based adsorption and exchange processes which
moderate the supply of major cations and anions to surface waters and
strongly influence the response of surface water ANC to changes in
deposition of nitrogen and sulfur.
3. Uncertainties in Modeled Processes
As discussed in the PA, from an uncertainty perspective, gaps in
field measurement data are related to uncertainties in modeled
processes and in the specific application of such models. As noted
above, processes that are embodied in an AAI-based standard are modeled
using the CMAQ atmospheric model and steady state ecological models.
These models are characterized in the ISA as being well established and
they have undergone extensive peer review. Nonetheless, the application
of these models for purposes of specifying the factors in the AAI
equation, on an ecoregion scale, is a new application that introduces
uncertainties, as noted below, especially in areas with limited
observational data that can be used to evaluate this specific
application. Understanding uncertainties in relevant modeled process
thus involves consideration of the uncertainties associated with
applying each model as well as the combination of these uncertainties
as the models are applied in combination within the AAI framework.
With regard to the application of CMAQ for purposes of use in an
AAI-based standard, the modeling of dry deposition has been identified
as having a relatively high degree of uncertainty. Due to a combination
of system complexity and resource constraints, there is no routine
observational basis for directly comparing modeled dry deposition and
measurements. Periodic dry deposition flux experiments covering a
variety of vegetation, surfaces and meteorology across seasons would

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enable a more robust evaluation of modeled deposition of nitrogen and
sulfur. Given the difficulty in acquiring dry deposition observations,
it becomes especially important to evaluate the model's ability to
capture temporal and spatial ambient air patterns of individual
nitrogen and sulfur species which are used to drive dry deposition
calculations in models. For example, reducing a generally acknowledged
positive bias in model-predicted SO2 relative to
observations is especially relevant to the AAI-based standard, as
SO2 deposition is a dominant contributor to total acidifying
deposition in the eastern U.S. With respect to oxidized nitrogen,
observations of individual NOy species are important as air
quality models calculate the individual deposition of each species. The
modeled transference ratios, TNOy and TSOx used
in factors F3 and F4 rely on CMAQ's ability to characterize both
deposition and concentration. Consequently, a better understanding of
the variability of these factors within and across ecoregions could be
achieved by improved availability of measured ambient concentrations
and deposition observations.
Steady state biogeochemical ecosystem modeling is used to develop
critical load estimates that are incorporated in the AAI equation
through factor F1. Consequently, the PA notes that an estimate of the
temporal response of surface water ANC to deposition and air
concentration changes is not directly available. Lacking a predicted
temporal response impairs the ability to conduct accountability
assessments down to the effects level. Accountability assessments would
examine the response of each step in the emissions source through air
concentration--deposition--surface water quality--biota continuum. The
steady state assumption at the ecosystem level does not impair
accountability assessments through the air concentration/deposition
range of that continuum. However, in using steady state ecosystem
modeling, several assumptions are made relative to the long-term
importance of processes related to soil adsorption of major ions and
ecosystem nitrogen dynamics. Because these models often were developed
and applied in glaciated areas with relatively thin and organically
rich soils, their applicability is relatively more uncertain in areas
such as those in the non-glaciated clay-based soil regions of the
central Appalachians. Consequently, it is desirable to develop the
information bases to drive simple dynamic ecosystem models that
incorporate more detailed treatment of subsurface processes, such as
adsorption and exchange processes and sulfate absorption.
4. Applying Knowledge of Uncertainties
An understanding of the relative uncertainties in a system assists
in setting priorities for data collection efforts and research, with
the expectation that such efforts would reduce uncertainties over time
and afford greater confidence in applications of an AAI-based standard.
Because of the uniquely wide breadth of pollutants and environmental
media addressed by an AAI-based multi-pollutant standard, there are a
wide range of uncertainties that are important to consider relative to
single pollutant standards that typically address only direct effects
of ambient air exposures. For an AAI-based standard, a reduction of the
uncertainties across the various modeled processes at the ecoregion
scale would lead to greater confidence in the degree of protection
afforded by the standard.
The PA notes that there is generally low uncertainty with regard to
the conceptual development and related major components of this
standard. In recognizing the scientific soundness of the basic
structure of this standard, the PA notes that future efforts would be
appropriately directed at expanding the availability of relevant data
for ecoregion-specific evaluation and application of the relevant
modeling of ecological and atmospheric processes, as identified above.
Such efforts would further support consideration of an AAI-based
standard and would guide field studies and analyses designed to improve
the longer-term confidence in such a standard.

G. CASAC Advice

The CASAC has advised EPA concerning the ISA, the REA, and the PA.
The CASAC has endorsed EPA's interpretation of the science embodied in
the ISA and the assessment approaches and conclusions incorporated in
the REA.
Most recently, CASAC has considered the information in the final PA
in providing its recommendations on the review of the new multi-
pollutant standard developed in that document and discussed above
(Russell and Samet, 2011a). In so doing, CASAC has expressed general
support for the conceptual framework of the standard based on the
underlying scientific information, as well as for the conclusions in
the PA with regard to indicators, form, averaging time, and level of
the standard that are appropriate for consideration by the Agency in
reaching decisions on the review of the secondary NAAQS for oxides of
nitrogen and sulfur:

The final Policy Assessment clearly sets out the basis for the
recommended ranges for each of the four elements (indicator,
averaging time, level and form) of a potential NAAQS that uses
ambient air indicators to address the combined effects of oxides of
nitrogen and oxides of sulfur on aquatic ecosystems, primarily
streams and lakes. As requested in our previous letters, the Policy
Assessment also describes the implications of choosing specific
combinations of elements and provides numerous maps and tabular
estimates of the spatial extent and degree of severity of NAAQS
exceedances expected to result from possible combinations of the
elements of the standard.
We believe this final PA is appropriate for use in determining a
secondary standard to help protect aquatic ecosystems from
acidifying deposition of oxides of sulfur and nitrogen. EPA staff
has done a commendable job developing the innovative Aquatic
Acidification Index (AAI), which provides a framework for a national
standard based on ambient concentrations that also takes into
account regional differences in sensitivities of ecosystems across
the country to effects of acidifying deposition. (Russell and Samet,
2011a)

The CASAC also recommended that as EPA moves forward in the
regulatory process ``some attention should be given to our residual
concern that the available data may reflect the more sensitive water
bodies and thus, the selection of percentiles of waterbodies to be
protected could be conservatively biased'' (Russell and Samet, 2011a).
In addition, CASAC found some improvements could be made to the
uncertainty analysis, as noted below. With respect to indicators, CASAC
supports the use of SOx and NOy as ambient air
indicators (discussed above in section III.A) and ANC as the ecological
indicator (discussed above in section III.B.1):
The use of NOy and SOx as atmospheric
indicators of oxides of nitrogen and sulfur atmospheric concentrations
is well justified. The use in the AAI of NOy and
SOx as atmospheric indicators of oxides of nitrogen and
sulfur concentrations is useful and corresponds with other efforts by
EPA. As we have stated previously, CASAC also agrees that ANC is the
most appropriate ecological indicator of aquatic ecosystem response and
resiliency to acidification (Russell and Samet, 2011a).
With respect to the form of the standard (discussed above in
section III.B), CASAC stated the following:

EPA has developed the AAI, an innovative ``form'' of the NAAQS
itself that incorporates

[[Page 46133]]

the multi-pollutant, multi-media, environmentally modified,
geographically variable nature of SOx/NOy
deposition-related aquatic acidification effects. With the caveats
noted below, CASAC believes that this form of the NAAQS as described
in the final Policy Assessment is consistent with and directly
reflective of current scientific understanding of effects of
acidifying deposition on aquatic ecosystems. (Russell and Samet,
2011a)
CASAC agrees that the spatial components of the form in the
Policy Assessment are reasonable and that use of Omernick's
ecoregions (Level III) is appropriate for a secondary NAAQs intended
to protect the aquatic environment from acidification * * * (Russell
and Samet, 2011a)

The ``caveats'' noted by CASAC include a recognition of the
importance of continuing to evaluate the performance of the CMAQ and
ecological models to account for model uncertainties and to make the
model-dependent factors in the AAI more transparent. In addition, CASAC
noted that the role of DOC and its effects on ANC would benefit from
further refinement and clarification (Russell and Samet, 2011a). While
CASAC expressed the view that the ``division of ecoregions into
`sensitive' and `non-sensitive' subsets, with a more protective
percentile applied to the sensitive areas, also seems reasonable''
(Russell and Samet, 2011a), CASAC also noted that there was the need
for greater clarity in specifying how appropriate screening criteria
would be applied in assigning ecoregions to these categories. Further,
CASAC identified potential biases in critical load calculations and in
the regional representativeness of available water chemistry data,
leading to the observation that a given percentile of the distribution
of estimated critical loads may be protective of a higher percentage of
surface waters in some regions (Russell and Samet, 2011a).
With respect to averaging time (discussed above in section III.C),
CASAC stated the following:

Considering the cumulative nature of the long-term adverse
ecological effects and the year-to-year variability of atmospheric
conditions (mainly in the amount of precipitation), CASAC concurs
with EPA that an averaging time of three to five years for the AAI
parameters is appropriate. A longer averaging time would mask
possible trends of AAI, while a shorter averaging time would make
the AAI being more influenced by the conditions of the particular
years selected. (Russell and Samet, 2011a)

With respect to level as well as the combination of level and form
as they are presented as alternative standards (discussed above in
sections III.D-E), CASAC stated the following:

CASAC agrees with EPA staff's recommendation that the ``level''
of the alternative AAI standards should be within the range of 20
and 75 [mu]eq/L. We also recognize that both the ``level'' and the
form of any AAI standard are so closely linked in their
effectiveness that these two elements should be considered together.
(Russell and Samet, 2011a)
When considered in isolation, it is difficult to evaluate the
logic or implications of selecting from percentiles (70th to 90th)
of the distribution of estimated critical loads for lakes in
sensitive ecoregions to determine an acceptable amount of deposition
for a given ecoregion. However, when these percentile ranges are
combined with alternative levels within the staff-recommended ANC
range of 20 to 75 microequivalents per liter ([mu]eq/L), the results
using the AAI point to the ecoregions across the country that would
be expected to require additional protection from acidifying
deposition. Reasonable choices were made in developing the form. The
number of acid sensitive regions not likely to meet the standard
will be affected both by choice of ANC level and the percentile of
the distribution of critical loads for lakes to meet alternative ANC
levels in each region. These combined recommendations provide the
Administrator with a broad but reasonable range of minimally to
substantially protective options for the standard. (Russell and
Samet, 2011a)

CASAC also commented on EPA's uncertainty analysis, and provided
advice on areas requiring further clarification in the proposed rule
and future research. The CASAC found it ``difficult to judge the
adequacy of the uncertainty analysis performed by EPA because of lack
of details on data inputs and the methodology used, and lack of clarity
in presentation'' (Russell and Samet, 2011a). In particular, CASAC
identified the need for more thorough model evaluations of critical
load and atmospheric modeling, recognizing the important role of models
as they are incorporated in the form of the standard. In light of the
innovative nature of the standard developed in the PA, CASAC identified
``a number of areas that should be the focus of further research''
(Russell and Samet, 2011a). While CASAC recognized that EPA staff was
able to address some of the issues in the PA, they also noted areas
``that would benefit from further study or consideration in potential
revisions or modifications to the form of the standard.'' Such research
areas include ``sulfur retention and mobilization in the soils,
aluminum availability, soil versus water acidification and ecosystem
recovery times.'' Further, CASAC encouraged future efforts to monitor
individual ambient nitrogen species, which would help inform further
CMAQ evaluations and the specification of model-derived elements in the
AAI equation (Russell and Samet, 2011a).

H. Administrator's Proposed Conclusions

Having concluded that the existing NO2 and
SO2 secondary standards are neither sufficiently protective
nor appropriate to address deposition-related effects associated with
oxides of nitrogen and sulfur (section II.D above), the Administrator
has considered whether it is appropriate at this time to set a new
multi-pollutant standard for that purpose, with a structure that would
better reflect the available science regarding acid deposition. In
considering this, she recognizes that such an appropriate standard, for
purposes of section 109(b) and (d) of the CAA,\15\ must in her judgment
be requisite to protect public welfare, such that it would be neither
more nor less stringent that necessary for that purpose. In particular,
she has focused on the new standard developed in the PA and reviewed by
CASAC, as discussed above. In so doing, the Administrator first
considered the extent to which there is a scientific basis for
development of such a standard, specifically with regard to a standard
that would provide protection from deposition-related aquatic
acidification in sensitive aquatic ecosystems in areas across the
country. As discussed above, the Administrator notes that the ISA
concludes that the available scientific evidence is sufficient to infer
a causal relationship between acidifying deposition of nitrogen and
sulfur in aquatic ecosystems, and that the deposition of oxides of
nitrogen and sulfur both cause such acidification under current
conditions in the U.S. Further, the ISA concludes that there are well-
established water quality and biological indicators of aquatic
acidification as well as well-established models that address
deposition, water quality, and effects on ecosystem biota, and that
ecosystem sensitivity to acidification varies across the country
according to present and historic nitrogen and sulfur deposition as
well as geologic, soil, vegetative, and hydrologic factors. Based on
these considerations, the Administrator agrees with the conclusion in
the PA, and supported by CASAC, that there is a strong scientific basis
for development

[[Page 46134]]

of a standard with the general structure presented in the PA.
---------------------------------------------------------------------------

\15\ Section 109(d)(1) requires that ``* * * the Administrator
shall complete a thorough review * * * and shall make such revisions
in such criteria and standards and promulgate such new standards as
may be appropriate under * * * subsection 109(b) of this section.''
[emphasis added]
---------------------------------------------------------------------------

The Administrator also recognizes that the conceptual framework for
an ecologically relevant, multi-pollutant standard, which was initially
explored in the REA and further developed in the PA, builds on the
information in the ISA. She notes that the structure of the standard
addresses the combined effects of deposition from oxides of nitrogen
and sulfur by characterizing the linkages between ambient
concentrations, deposition, and aquatic acidification, and that the
structure of the standard takes into account relevant variations in
these linkages across the country. She recognizes that while the
standard is innovative and unique, the structure of the standard is
well grounded in the science underlying the relationships between
ambient concentrations of oxides of nitrogen and sulfur and the aquatic
acidification related to deposition of nitrogen and sulfur associated
with such ambient concentrations.
While the Administrator recognizes the strong scientific foundation
for the structure of an AAI-based standard, she also recognizes that
the standard depends on atmospheric and ecological modeling, based on
appropriate data, to specify the terms of an equation that incorporates
the linkages between ambient concentrations, deposition, and aquatic
acidification. This equation, which defines an aquatic acidification
index (AAI), has the effect of translating spatially variable
ecological effects into a potential national standard. With respect to
establishing the specific terms of this equation, there are a number of
inherent uncertainties and complexities that are relevant to the
question of whether it is appropriate under section 109 to set a
specific AAI-based standard at this time, recognizing that such a
standard must be requisite to protect public welfare without being
either more or less stringent than necessary for this purpose. As
discussed above, these uncertainties and complexities generally relate
not to the structure of the standard, but to the quantification of the
various elements of the standard, such as the F factors discussed
earlier in this section and their representativeness at an ecoregion
scale. These uncertainties and complexities currently limit efforts to
characterize the degree of protectiveness that would be afforded by
such a standard, within the ranges of levels and forms identified in
the PA, and the representativeness of F factors in the AAI equation
described above and in the PA. These important uncertainties have been
generally categorized as limitations in available field data as well as
uncertainties that are related to reliance on the application of
ecological and atmospheric modeling at the ecoregion scale to specify
the various elements of the AAI.
With regard to data limitations, the Administrator observes that
there are several important limitations in the available data upon
which elements of the AAI are based. For example, while ambient
measurements of NOy are made as part of a national
monitoring network, the monitors are not located in locations that are
representative of sensitive aquatic ecosystems. While air and water
quality data are generally available in areas in the eastern U.S.,
there is relatively sparse coverage in mountainous western areas where
a number of sensitive aquatic ecosystems are located. Further, even in
areas where relevant data are available, small sample sizes impede
efforts to characterize the representativeness of the available data,
which was noted by CASAC as being of particular concern. Also,
measurements of reduced forms of nitrogen are available from only a
small number of monitoring sites, and emission inventories for reduced
forms of nitrogen used in atmospheric modeling are subject to
considerable uncertainty.
With regard to uncertainties related to the use of ecological and
atmospheric modeling, the Administrator notes in particular that model
results are difficult to evaluate due to a lack of relevant
observational data. For example, relatively large uncertainties are
introduced by a lack of data with regard to pre-industrial
environmental conditions and other parameters that are necessary inputs
to critical load models that are the basis for factor F1 in the AAI
equation. Also, observational data are not generally available to
evaluate the modeled relationships between nitrogen and sulfur in the
ambient air and associated deposition, which are the basis for the
other factors (i.e., F2, F3, and F4) in the AAI equation.
In combination, these limitations and uncertainties result in a
considerable degree of uncertainty as to how well the quantified
elements of the AAI standard would predict the actual relationship
between varying ambient concentrations of oxides of nitrogen and sulfur
and steady state ANC levels across the distribution of water bodies
within the various ecoregions in the U.S. Because of this, there is
considerable uncertainty as to the actual degree of protectiveness that
such a standard would provide, especially for acid-sensitive
ecoregions. The Administrator recognizes that the AAI equation, with
factors quantified in the ranges discussed above and described more
fully in the PA, generally performs well in identifying areas of the
country that are sensitive to such acidifying deposition and indicates,
as expected, that lower ambient levels of oxides of nitrogen and sulfur
would lead to higher calculated AAI values. However, the uncertainties
discussed here are critical for determining the actual degree of
protection that would be afforded such areas by any specific target ANC
level and percentile of water bodies that would be chosen in setting a
new AAI-based standard, and thus for determining an appropriate AAI-
based standard that meets the requirements of section 109.
In considering these uncertainties, the Administrator notes that
CASAC acknowledged that important uncertainties remain that would
benefit from further study and data collection efforts, which might
lead to potential revisions or modifications to the form of the
standard developed in the PA. She also notes that CASAC encouraged the
Agency to engage in future monitoring and model evaluation efforts to
help inform the specification of model-derived elements in the AAI
equation.
Based on the above considerations, the Administrator has determined
that it is not appropriate under section 109 to set a new multi-
pollutant standard to address deposition-related effects of oxides of
nitrogen and sulfur on aquatic acidification at this time. Setting a
NAAQS generally involves consideration of the degree of uncertainties
in the science and other information, such as gaps in the relevant data
and, in this case, limitations in the evaluation of the application of
relevant ecological and atmospheric models at an ecoregion scale. As
noted above, the issue here is not a question of uncertainties about
the scientific soundness of the structure of the AAI, but instead
uncertainties in the quantification and representativeness of the
elements of the AAI as they vary in ecoregions across the country. At
present, these uncertainties prevent an understanding of the degree of
protectiveness that would be afforded to various ecoregions across the
country by a new standard defined in terms of a specific nationwide
target ANC level and a specific percentile of water bodies for acid-
sensitive ecoregions and thus prevent identification of an appropriate
standard.. The Administrator has considered whether these uncertainties
could be appropriately accounted for by choosing either a more or less
protective target ANC level and percentile of water bodies than would
otherwise be chosen if the uncertainties did not substantially limit
the confidence that can

[[Page 46135]]

appropriately be ascribed to the quantification of the AAI elements.
However, in the Administrator's judgment, the uncertainties are of such
nature and magnitude that there is no reasoned way to choose such a
specific nationwide target ANC level or percentile of water bodies that
would appropriately account for the uncertainties, since neither the
direction nor the magnitude of change from the target level and
percentile that would otherwise be chosen can reasonably be ascertained
at this time.
Based on the above considerations, the Administrator judges that
the current limitations in relevant data and the uncertainties
associated with specifying the elements of the AAI based on modeled
factors are of such nature and degree as to prevent her from reaching a
reasoned decision such that she is adequately confident as to what
level and form (in terms of a selected percentile) of such a standard
would provide any particular intended degree of protection of public
welfare that the Administrator determined satisfied the requirements to
set an appropriate standard under section 109. While acknowledging that
CASAC supported moving forward to establish the standard developed in
the PA, the Administrator also observes that CASAC supported conducting
further field studies that would better inform the continued
development or modification of such a standard. Given the large
uncertainties and complexities inherent in quantifying the elements of
such a standard, largely deriving from the unprecedented nature of the
standard under consideration in this review, and having fully
considered CASAC's advice, the Administrator provisionally concludes
that it is premature to set a new, multi-pollutant secondary standard
for oxides of nitrogen and sulfur at this time, and as such she is
proposing not to set such a new secondary standard.
While it is premature to set such a multi-pollutant standard at
this time, the Administrator determines that the Agency should
undertake a field pilot program to gather additional data, and that it
is appropriate that such a program be undertaken before, rather than
after, reaching a decision to set such a standard. As described below
in section IV, the purpose of the program is to collect and analyze
data so as to enhance our understanding of the degree of protectiveness
that would likely be afforded by a standard based on the AAI as
developed in the PA. This will provide additional information to aid
the Agency in considering an appropriate multi-pollutant standard,
specifically with respect to the acidifying effects of deposition of
oxides of nitrogen and sulfur. PA. Data generated by this field program
will also support development of an appropriate monitoring network that
would work in concert with such a standard to result in the intended
degree of protection. The data and analyses generated as a result of
this program will serve to inform the next review of the NAAQS for
oxides of nitrogen and sulfur. The information generated during the
field program can also be used to help state agencies and EPA better
understand how an AAI-based standard would work in terms of the
implementation of such a standard.
Based on the above considerations, the Administrator is proposing
not to set a new multi-pollutant AAI-based secondary standard for
oxides of nitrogen and sulfur in this review. In reaching this
decision, the Administrator recognizes that the new NO2 and
SO2 primary 1-hour standards set in 2010, while not
ecologically relevant for a secondary standard, will nonetheless result
in reductions in oxides of nitrogen and sulfur that will directionally
benefit the environment by reducing NOy and SOx deposition to sensitive
ecosystems. EPA is proposing to revise the secondary standards by
adding secondary standards identical to the NO2 and
SO2 primary 1-hour standards set in 2010. More specifically,
EPA is proposing a 1-hour secondary NO2 standard set at a
level of 100 ppb and a 1-hour secondary SO2 standard set at
a level of 75 ppb. While this will not add secondary standards of an
ecologically relevant form to address deposition-related effects, it
will directionally provide some degree of additional protection. This
is consistent with the view that the current secondary standards are
neither sufficiently protective nor appropriate in form, but that it is
not appropriate to propose to set a new, ecologically relevant multi-
pollutant secondary standard at this time, for all of the reasons
discussed above.
While not a basis for this decision, the Administrator also
recognizes that a new, innovative AAI-based standard would raise
significant implementation issues that would need to be addressed
consistent with the CAA requirements for implementation-related actions
following the setting of a new NAAQS. It will take time to address
these issues, during which the Agency will be conducting a field pilot
program to gather relevant data and the environment will benefit from
reductions in oxides of nitrogen and sulfur resulting from the new
NO2 and SO2 primary standards, as noted above, as
well as reductions expected to be achieved from EPA's Cross-State Air
Pollution Rule and Mercury and Air Toxics standards. These
implementation-related issues are discussed in more detail below in
section IV.A.5.
The Administrator solicits comment on all aspects of this proposed
decision, including the framework and elements of a multi-pollutant
standard for oxides of nitrogen and sulfur to address deposition-
related effects on sensitive ecosystems, with a focus on aquatic
acidification, and the uncertainties and complexities associated with
the development of such a standard at this time. The Administrator also
solicits comment on the field pilot program and related monitoring
methods as discussed below in section IV.

IV. Field Pilot Program and Ambient Monitoring

This section describes EPA's plans for a field pilot program and
the evaluation of monitoring methods for ambient air indicators of
NOy and SOx to implement the Administrator's
decision to undertake such a field monitoring program in conjunction
with her decision to propose not to set a new multi-pollutant secondary
standard in this review, as discussed above in section III.H. As noted
above and discussed below in section IV.A, the field pilot program is
intended to collect and analyze data so as to enhance our understanding
of the degree of protectiveness that would likely be afforded by a
standard based on the AAI as developed in the PA. Data generated by
this field program would also support development of an appropriate
monitoring network that would work in concert with such a standard to
result in the intended degree of protection. As discussed below in
section IV.B, the evaluation of monitoring methods focuses on the
development of Federal Reference Methods/Federal Equivalent Methods
(FRM/FEM) for NOy and SOx. The EPA notes that the
monitoring program described here is intended to be coordinated with
EPA's CASTNET as a supplement to existing monitoring programs and is
beyond the scope of the current CASTNET program.

A. Field Pilot Program

This section presents the objectives of a field pilot program
(section IV.A.1) that would gather relevant field data over a 5-year
period in a sample of three to five sensitive ecoregions across the
country. An overview of the scope and structure of the field program,
with a focus on measurements of ambient air indicators of oxides of
nitrogen and

[[Page 46136]]

sulfur, is presented in section IV.A.2. Section IV.A.3 explains the
role of additional complementary measurements beyond the ambient air
indicators that would be included in the program, and section IV.A.4
discusses a parallel longer-term research agenda, both of which are
guided by the uncertainties discussed above in section III. Section
IV.A.5 identifies implementation challenges presented by an AAI-based
standard that could be addressed in parallel with a field pilot
program. Section IV.A.6 discusses engagement with stakeholder groups as
part of the planned pilot program.
1. Objectives
Consideration of a new multi-pollutant standard to address
deposition-related effects on sensitive aquatic ecoregions raises
unique challenges relative to those typically raised in reviews of
existing NAAQS for which an established network of FRM/FEM monitors,
designed to measure the indicator pollutant, is generally available.
The primary goal of this field pilot program, and the related
monitoring program discussed in section IV.B, is to enhance our
understanding of the degree of protectiveness that would likely be
afforded by a standard based on the AAI, as described above in section
III, so as to aid the Agency in considering an appropriate multi-
pollutant standard that would be requisite to protect public welfare
consistent with section 109 of the CAA, through the following
objectives:
(1) Evaluate measurement methods for the ambient air indicators of
NOy and SOx and consider designation of such
methods as FRMs;
(2) Examine the variability and improve characterization of
concentration and deposition patterns of NOy and
SOx, as well as reduced forms of nitrogen, within and across
a number of sensitive ecoregions across the country;
(3) Develop updated ecoregion-specific factors (i.e., F1 through
F4) for the AAI equation based in part on new observed air quality data
within the sample ecoregions as well as on updated nationwide air
quality model results and expanded critical load data bases, and
explore alternative approaches for developing such representative
factors;
(4) Calculate ecoregion-specific AAI values using observed
NOy and SOx data and updated ecoregion-specific
factors to examine the extent to which the sample ecoregions would meet
a set of alternative AAI-based standards;
(5) Develop air monitoring network design criteria for an AAI-based
standard;
(6) assess the use of total nitrate measurements as a potential
alternative indicator for NOy;
(7) Support related longer-term research efforts, including
enhancements to and evaluation of modeled dry deposition algorithms;
and
(8) Facilitate stakeholder engagement in addressing implementation
issues associated with possible future adoption of an AAI-based
standard.
2. Overview of Field Pilot Program
The CASTNET program (Figure IV-1) affords an available
infrastructure relevant to an AAI-based standard, given the location of
sites in some acid-sensitive ecoregions and various measurements of
sulfur and nitrogen species. The EPA plans to use CASTNET sites in
selected acid-sensitive ecoregions to serve as the platform for this
pilot program, potentially starting in late 2012 and extending through
2018. The CASTNET sites in three to five ecoregions in acid-sensitive
areas would collect NOy and SOx (i.e.,
SO2 and p-SO4) measurements over a 5-year period.
The initial step in developing a data base of observed ambient air
indicators for oxides of nitrogen and sulfur requires the addition of
NOy samplers at the pilot study sites so that a full
complement of indicator measurements are available to calculate AAI
values. These CASTNET sites would also be used to make supplemental
observations useful for evaluation of CMAQ's characterization of
factors F2 -F4 in the AAI equation.
The selected ecoregions would account for geographic variability by
including regions from across the U.S., including the east, upper
midwest and west. Each selected region would have at least two existing
CASTNET sites. Each of the pilot CASTNET sites would be used to
evaluate the performance of the established methods, data retrieval and
reporting procedures used in the AAI equation.

[[Page 46137]]

[GRAPHIC] [TIFF OMITTED] TP01AU11.030

Over the course of this 5-year pilot program, the most current
national air quality modeling, based on the most current national
emissions inventory, would be used to develop an updated set of F2--F4
factors. A parallel multi-agency national critical load data base
development effort would be used as the basis for calculating updated
F1 factors. As discussed above in section III.B, these factors would be
based on average parameter values across an ecoregion. Using this new
set of F factors, observations of NOy and SOx
derived from the pilot program, averaged across each ecoregion, would
be used to calculate AAI values in the sample ecoregions. The data from
the pilot program would also be used to examine alternative approaches
to generating representative air quality values, such as examining the
appropriateness of spatial averaging in areas of high spatial
variability.
3. Complementary Measurements
Complementary measurements may be performed at some sites in the
pilot network to reduce uncertainties in the recommended methods and
better characterize model performance and application to the AAI. The
CASAC Air Monitoring and Methods Subcommittee (AMMS) advised EPA that
such supplemental measurements were of critical importance in a field
measurement program related to an AAI-based standard (Russell and
Samet, 2011b).
Candidate complementary measurements to address sulfur, in addition
to those provided by the CASTNET filter pack (CFP), include trace gas
continuous SO2 and speciated PM2.5 measurements.
The co-located deployment of a continuous SO2 analyzer with
the CFP for SO2 will provide test data for determining
suitability of continuous SO2 measurements as a Federal
Equivalent Method (FEM), as well as producing valuable time series data
for model evaluation purposes. The weekly averaging time provided by
the CFP adequately addresses the annual-average basis of an AAI-based
secondary standard, but would not be applicable to short-term (i.e., 1-
hour) averages associated with the primary SO2 standard.
Conversely, because of the low concentrations associated with many
acid-sensitive ecoregions, existing SO2 Federal Reference
Methods (FRMs) designated for use in determining compliance with the
primary standard would not necessarily be appropriate for use in
conjunction with an AAI-based secondary standard.
Co-locating the PM2.5 sampler used in the EPA Chemical
Speciation Network and the Interagency Monitoring of Protected Visual
Environments (IMPROVE) network at pilot network sites would allow for
characterizing the relationship between the CFP-derived p-
SO4 and the speciation samplers used throughout the state
and local air quality networks. Note that CASTNET already has several
co-located IMPROVE chemical speciation samplers. Because the AAI
equation is based on concentration of p-SO4, the original
motivation for capturing all particle size fractions is not as
important relative to simply capturing the concentration of total p-
SO4.
Candidate measurements to complement oxidized nitrogen
measurements, in addition to the CFP, include a mix of continuous and
periodic sampling for the dominant NOy species, namely NO,
true NO2, PAN,

[[Page 46138]]

HNO3, and p-NO3. While there are several
approaches to acquiring these measurements, perhaps the most efficient
strategy would take advantage of the available CFP for total nitrate,
and add a three-channel chemiluminescence instrument that will cycle
between NOy, true NO2 and NO by adding photolytic
detection for true NO2. Other options for measuring true
NO2 would include adding either a stand-alone photolytic or
cavity ring-down spectroscopy instrument. Measurements of PAN may be
acquired either on a periodic basis through canister sampling and
subsequent laboratory analysis or through emerging in-situ sampling and
analysis methods. Although the CFP yields a reliable measurement of
total nitrate, the t-NO3 (i.e., the sum of HNO3
and p-NH4) value, strong consideration may be given to
direct measurement of HNO3, which has the highest deposition
velocity of all the dominant NOy species. Similar to the use
of continuous SO2 data, these speciated NOy data
serve two purposes: evaluating total NOy instrument behavior
and evaluating air quality models. The measurement of individual
NOy species can be used to generate site-specific
NOy values for comparison to modeled NOy, and
will likely provide insight into and improvement of modeled dry
deposition.
The CASAC AMMS (Russell and Samet, 2011b) recommended that EPA
consider the use of t-NO3 obtained from CASTNET sampling as
an indicator for NOy, reasoning that t-NO3 is
typically a significant fraction of deposited oxidized nitrogen in
rural environments and CASTNET measurements are widely available.
Collection of this data would support further consideration of using
the CFP for t-NO3 as the indicator of oxides of nitrogen for
use in an AAI-based secondary standard.
The CASAC AMMS also recommended that total NHx
(NH3 and p-NH4) be considered as a proxy for
reduced nitrogen species, reasoning that the subsequent partitioning to
NH3 and p-NH4 may be estimated using equilibrium
chemistry calculations. Reduced nitrogen measurements are used to
evaluate air quality modeling which is used in generating factor F2.
Additional studies are needed to determine the applicability of
NHx measurements and calculated values of NH3 and
NH4 to the AAI.
The additional supplemental measurements of speciated
NOy, continuous SO2 and NHx will be
used in future air quality modeling evaluation efforts. Because there
often is significant lag in the availability of contemporary emissions
data to drive air quality modeling, the complete use of these data sets
will extend beyond the 5-year collection period of the pilot program.
Consequently, the immediate application of those data will address
instrument performance comparisons that explore the feasibility of
using continuous SO2 instruments in rural environments, and
using the speciated NOy data to assess NOy
instrument performance. Although contemporary air quality modeling will
lag behind measurement data availability, the observations can be used
in deposition models to compare observed transference ratios with the
previously calculated transference ratios to test temporal stability of
the ratios.
An extended water quality sampling effort should parallel the air
quality measurement program to address some of the uncertainties
related to factor F1 and the representativeness of the nth percentile
critical load as discussed in section III.B.5.b.i. The objective of the
water quality sampling would be to develop a larger data base of
critical loads in each of the pilot ecoregions such that the nth
percentile can adequately be characterized in terms of representing all
water bodies. Opportunities to leverage and perhaps enhance existing
ecosystem modeling efforts enabling more advanced critical load
modeling and improved methods to estimate base cation production would
be pursued. For example, areas with ongoing research studies producing
data for dynamic critical load modeling would be considered when
selecting the pilot ecoregions.
4. Complementary areas of research
The EPA recognizes that a source of uncertainty in an AAI-based
secondary standard that would not be directly addressed in the pilot
program stems from the uncertainty in the model used to link
atmospheric concentrations to dry deposition fluxes. Currently, there
are no ongoing direct dry deposition measurement studies at CASTNET
sites that can be used to evaluate modeled results. It was strongly
recommended by CASAC AMMS that a comprehensive sampling-intensive study
be conducted in at least one, preferably two sites in different
ecoregions to assess characterization of dry deposition of sulfur and
nitrogen. These sites would be the same as those for the complementary
measurements described above, but they would afford an opportunity to
also complement dry deposition process research that benefits from the
ambient air measurements collected in the pilot program. The concerns
regarding uncertainties underlying an AAI-based secondary standard
suggest that research that includes dry deposition measurements and
evaluation of dry deposition models should be a high priority.
Similar leveraging should be pursued with respect to ecosystem
research activities. For example, studies that capture a suite of soil,
vegetation, hydrological, and water quality properties that can help
evaluate more advanced critical load models would complement the
atmospheric-based pilot program. In concept, such studies could provide
the infrastructure for true multi-pollutant, multi-media ``super''
sites assuming the planning, coordination, and resource facets can be
aligned. While this discussion emphasizes the opportunity of leveraging
ongoing research efforts, consideration could be given to explicitly
including related research components directly in the pilot program.
5. Implementation challenges
The CAA requires that once a NAAQS is established, designation and
implementation must move forward. With a standard as innovative as the
AAI-based standard considered in this review, the Administrator
believes that its success will be greatly improved if, while additional
data are being collected to reduce the uncertainties discussed above,
the implementing agencies and other stakeholders have an opportunity to
discuss and thoroughly understand how such a standard would work. And
since, as noted above, emissions reductions that are directionally
correct to reduce aquatic acidification will be occurring as a result
of other CAA programs, the Administrator believes that this period of
further discussion will not delay progress but will ensure that once
implementation is triggered, agencies will be prepared to implement it
successfully.
Consideration of an AAI-based secondary standard for oxides of
nitrogen and sulfur would present significant implementation challenges
because it involves multiple, regionally-dispersed pollutants and
relatively complex compliance determinations based on regionally
variable levels of NOy and SOx concentrations
that would be necessary to achieve a national ANC target. The
anticipated implementation challenges fall into three main categories:
monitoring and compliance determinations for area designations, pre-
construction permit application analyses of individual source impacts,
and State Implementation Plan (SIP) development. Several overarching
implementation questions that we

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anticipate will be addressed in parallel with the field pilot program's
five-year data collection period include:
(1) What are the appropriate monitoring network density and siting
requirements to support a compliance system based on ecoregions?
(2) Given the unique spatial nature of the secondary standard
(e.g., ecoregions), what are the appropriate parameters for
establishing nonattainment areas?
(3) How can new or modified major sources of oxides of nitrogen and
oxides of sulfur emissions assess their ambient impacts on the standard
and demonstrate that they are not causing or contributing to a
violation of the NAAQS for preconstruction permitting? To what extent
does the fact that a single source may be impacting multiple areas,
with different acid sensitivities and variable levels of NOy
and SOx concentrations that would be necessary to achieve a
national ANC target, complicate this assessment and how can these
additional complexities best be addressed?
(4) What additional tools, information, and planning structures are
needed to assist states with SIP development, including the assessment
of interstate pollutant transport and deposition?
(5) Would transportation conformity apply in nonattainment and
maintenance areas for this secondary standard, and, if it does, would
satisfying requirements that apply for related primary standards (e.g.,
ozone, PM2.5, and NO2) be demonstrated to satisfy
requirements for this secondary standard?
6. Final Monitoring Plan Development and Stakeholder Participation
The existing CASTNET sampling site infrastructure provides an
effective means of quickly and efficiently deploying a monitoring
program to support potential implementation of an AAI-based secondary
standard, and also provides an additional opportunity for federally
managed networks to collaborate and support the states, local agencies
and tribes (SLT) in determining compliance with a secondary standard. A
collaborative effort would help to optimize limited federal and SLT
monitoring funds and would be beneficial to all involved. The CASTNET
is already a stakeholder-based program with over 20 participants and
contributors, including federal, state and tribal partners.
The CASAC AMMS generally endorsed the technical approaches used in
CASTNET, but concerns were raised by individual representatives of
state agencies concerning the perception of EPA-controlled management
aspects of CASTNET and data ownership. Potential approaches to resolve
these issues will be developed and evaluated in existing National
Association of Clean Air Agencies (NACAA)/EPA ambient air monitoring
workgroups. The EPA Office of Air and Radiation (which includes the
Office of Air Quality Planning Standards, OAQPS; and the Office of
Atmospheric Program's Clean Air Markets Division, OAP-CAMD), and their
partners on the NACAA monitor steering committee will develop a
prioritized specific plan that identifies the three to five ecoregions
and the instrumentation to be deployed. The EPA anticipates that a cost
estimate of the plan with priorities and options will be developed by
January, 2012. Although this pilot program is focused on data
collection, the plan will include details of the data analysis
approaches as well as a vehicle that incorporates engagement from those
within EPA and SLTs to foster progress on the implementation questions
noted above in section IV.A.5.
If an AAI-based secondary standard were to be set in the future,
deployment of a full national network would follow the pilot monitoring
program. The number of sites deployed in the network will lead to
increased confidence in capturing spatial patterns of air quality.
Recognizing that this section presents the general elements of the
field pilot programs, EPA intends to develop a more detailed field
pilot program plan through a process that will engage the air quality
management and research (atmospheric and ecosystem) communities, as
well as other federal agencies, state and local agencies, and non-
government based centers of expertise. The EPA is seeking comment and
input on all aspects of this field pilot program.

B. Evaluation of Monitoring Methods

The EPA generally relies on monitoring methods that have been
designated as FRMs or FEMs for the purpose of determining the
attainment status of areas with regard to existing NAAQS. Such FRMs or
FEMs are generally required to measure the air quality indicators that
are compared to the level of a standard to assess compliance with a
NAAQS. Prior to their designation by EPA as FRM/FEMs through a
rulemaking process, these methods must be determined to be applicable
for routine field use and need to have been experimentally validated by
meeting or exceeding specific accuracy, reproducibility, and
reliability criteria established by EPA for this purpose. As discussed
above in section III.A, the ambient air indicators being considered for
use in an AAI-based standard include SO2, particulate
sulfate (p-SO4), and total reactive oxides of nitrogen
(NOy).
The CASTNET provides a well established infrastructure that would
meet the basic location and measurement requirements of an AAI-based
secondary standard given the rural placement of sites in acid sensitive
areas. In addition, CFPs currently provide very economical weekly,
integrated average concentration measurements of SO2, p-
SO4, ammonium ion (NH4) and t-NO3, the
sum of HNO3 and p-NO3.
While routinely operated instruments that measure SO2,
p-SO4, NOy and/or t-NO3 exist,
instruments that measure p-SO4, NOy, t-
NO3, or the CFP for SO2 have not been designated
by EPA as FRMs or FEMs. The EPA's Office of Research and Development
has initiated work that will support future FRM designations by EPA for
SO2 and p-SO4 measurements based on the CFP. Such
a designation by EPA could be done for the purpose of facilitating
consistent research related to an AAI-based standard and/or in
conjunction with setting and supporting an AAI-based secondary
standard.
Based on extensive review of literature and available data, the EPA
has identified potential methods that appear suitable for measuring
each of the three components of the indicators. These three methods are
being considered as new FRMs to be used for measuring the ambient
concentrations of the three components that would be needed to
determine compliance with an AAI-based secondary standard.
For the SO2 and p-SO4 measurements, EPA is
considering the CFP method, which provides weekly average concentration
measurements for SO2 and p-SO4. This method has
been used in the EPA's CASTNET monitoring network for 15 years, and
strongly indicates that it will meet the requirements for use as an FRM
for the SO2 and p-SO4 concentrations for an AAI-
based secondary standard.
Although the CFP method would provide measurements of both the
SO2 and p-SO4 components in a unified sampling
and analysis procedure, individual FRMs will be considered for each.
The EPA recognizes that an existing FRM to measure SO2
concentrations using ultra-violet fluorescence (UVF) exists (40 CFR
Part 50, Appendix A-1) for the purpose of monitoring compliance for the
primary SO2 NAAQS. However, several factors suggest that the
CFP method would be

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superior to that UVF FRM for monitoring compliance with an AAI-based
secondary standard and will be discussed in more detail below.
For monitoring the NOy component, a continuous analyzer
for measuring NOy is commercially available and is
considered to be suitable for use as an FRM. This method is similar in
design to the existing NO2 FRM (described in 40 CFR Part 50,
Appendix F), which is based on the ozone chemiluminescence measurement
technique. The method is adapted to and further optimized to measure
all NOy. However, this NOy method requires
further evaluation before it can be fully confirmed as a suitable FRM.
The EPA is currently completing a full scientific assessment of the
NOy method to determine whether it would be appropriate to
consider for designation by EPA as an FRM. Specific details on these
three methods are given below.
On February 16, 2011, EPA presented this set of potential FRMs to
the CASAC AMMS for their consideration and comment. In response, the
CASAC AMMS stated that, overall, it believes that EPA's planned
evaluation of methods for measuring NOy, SO2 and
p-SO4 as ambient air indicators is a suitable approach in
concept. On supporting the CFP method as a potential FRM for
SO2, CASAC stated that they felt that the CFP is adequate
for measuring long-term average SO2 gas concentrations in
rural areas with low levels (less than 5 parts per billion by volume
(ppbv)) and is therefore suitable for consideration as an FRM. For p-
SO4, CASAC generally supports the use of the CFP as a
potential FRM for measuring p-SO4 for an AAI-based secondary
standard. The method has been relatively well-characterized and
evaluated, and it has a documented, long-term track record of
successful use in a field network designed to assess spatial patterns
and long-term trends.
On supporting the photometric NOy method as a potential
FRM, CASAC concluded that the existing NOy method is
generally an appropriate approach for the indicator. However, CASAC
agrees that additional characterization and research is needed to fully
understand the method in order to designate it as a FRM. The EPA is now
soliciting public comment on these methods as to their adequacy,
suitability, and relative merits as FRMs for purposes of monitoring to
determine compliance with an AAI-based secondary standard.
1. Potential FRMs for SO2 and p-SO4
The CFP is a combined, integrated sampling and analysis method
based on the well-established measurement technology that has been used
extensively in EPA's CASTNET monitoring network (see http://www.epa.gov/castnet). This method is in current use at over 80
monitoring sites and has been in use at not less than 40 sites for over
15 years. This method employs a relatively simple and inexpensive
sampler and uses four 47-mm filters placed in an open-faced filter pack
to simultaneously collect integrated filter samples for the
SO2 and p-SO4 components. In addition, the CFP is
also capable of the collection of t-NO3, the sum of
HNO3 and p-NO3.
The first stage of the filter pack assembly contains a Teflon[reg]
filter that collects p-SO4^2- and p-
NO3, the second stage contains a nylon filter that collects
SO2 (as SO4^2-) and HNO3,
and the third stage contains two cellulose fiber filters impregnated
with potassium carbonate (K2CO3) that collect any
remaining SO2 (as SO4^2-). The sampler
collects 1-week integrated samples at a very low, controlled flow rate
(1.5 or 3 L/min) in an attempt simulate actual deposition. Weekly
averaged SO2 and p-SO4 concentrations could then
be averaged over a 1-year period to calculate annual average values.
Upon sample completion, the species-specific filters are extracted,
with subsequent analysis by the well-established and documented ion
chromatographic (IC) analytical technique. During the IC analysis, an
aliquot of a filter extract is injected into a stream of eluent (ion
chromatography mobile phase, generally a millimolar-strength solution
of carbonate-bicarbonate) and passed through a series of ion
exchangers. The anions of interest are separated on the basis of their
relative affinities for a low capacity and the strongly basic anion
exchanger (guard and separator column). The separated anions are
directed onto a cation exchanger (suppressor column) where they are
converted to their highly conductive acid form, and the eluent is
converted to a weakly conductive form. The now-separated anions, each
in their acid form, are measured by conductivity. They are identified
on the basis of retention time compared to that of standards and
quantified by measurement of peak area compared to the peak areas of
calibration standards.
Calibration and quality assurance for the method are applied to the
sample filters, the analytical processes, and the flow rate measurement
and control aspects of the sampler. Overall method performance is
typically assessed with collocated samplers. These quality assurance
techniques are routinely used and have proved adequate for other types
of FRMs and equivalent methods in air monitoring network service.
The measurement and analytical procedures and past performance data
associated with the CFP method are well documented and available
through Quality Assurance Performance Plans (QAPPs), Standard Operating
Procedures (SOPs) and annual reports (US EPA, 2010a and 2010b). The
accumulated database on the CFP method is substantial and indicates
that the method is sound, stable and has good reliability in routine,
field operation. Data quality assessment results show the method to
have good reproducibility, with collocated and analytical precision
values in the range of 2 percent to 10 percent (excluding very low
concentration measurements near the method detection limits; US EPA
2010b).
Data quality objectives (DQOs) for a new FRM would be based upon
current DQOs being used for this method by EPA's OAP/CAMD and the NPS,
the federal managers of CASTNET (US EPA, 2010a). In its current state,
the CFP method is expected to meet or exceed (as past CASTNET data have
indicated; US EPA, 2010b) the expected FRM DQOs, even when deployed in
new monitoring networks outside of CASTNET. In addition, CASTNET
samples have agreed favorably with other measures of SO2 and
p-SO4 in comparison studies. For example, in direct
comparison with an annular denuder sampler (ADS) method, CASTNET/ADS
ratios for SO2 and p-SO4 were generally on the
order of 0.9-1.1 (Lavery et al, 2009; Sickles et al, 1999; Sickles et
al, 2008), thus illustrating the accuracy of the CFP method in the
determination of long-term average SO2 and p-SO4
concentrations. The EPA believes that the CFP method would be fully
adequate as an FRM in determining yearly average SO2 and p-
SO4 concentrations for compliance determination purposes.
The EPA recognizes that an existing FRM for SO2 has
proven adequate for the purposes of monitoring compliance for the
primary SO2 NAAQS, specifically the newly-promulgated 1-hour
standard. However, this FRM is better suited to the shorter-term,
higher concentration primary and secondary SO2 NAAQS, and
there is substantial uncertainty as to the adequacy of this
SO2 FRM for monitoring the lower concentrations relevant to
determining compliance with an AAI-based secondary standard. The
performance specifications for SO2 FRM analyzers (40 CFR
Part 53, Table B-1) require a lower detectable limit (LDL) of 0.002

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ppm for the standard measurement range and 0.001 ppm for the lower
measurement range. These requirements correspond to mass per unit
volume concentrations of 5.24 and 2.62 [micro]g/m\3\, respectively.
Analysis of 2009 CASTNET data shows that of the 84 CASTNET sampling
sites, 63 measured annual average SO2 concentrations below
even the lower of these LDL requirements of 2.62 [micro]g/m\3\ for the
lower range SO2 FRM (US EPA, 2010a). In addition, 11 of the
84 sites measured annual (2009) average SO2 concentrations
very near or below the manufacturers' reported detection limits for
trace level UVF SO2 monitors. Further, it is likely that the
number of sites with annual average SO2 concentration below
both the SO2 FRM LDL and the manufacturers reported
detection limits will increase due to expected declines in mean
SO2 concentrations (US EPA, 2010b). For these reasons, EPA
is considering the CFP method for use as the FRM for monitoring the
SO2 component of an ambient air indicator for oxides of
sulfur, with a recommendation for additional study and data collection
to evaluate further the possible applicability of the continuous UVF
SO2 FRM for this purpose.
2. Potential FRM for NOy
Atmospheric concentrations of NOy are measured
continuously by an analyzer that photometrically measures the light
intensity, at wavelengths greater than 600 nanometers (nm), resulting
from the chemiluminescent reaction of ozone (O3) with NO in
sampled air. This method is very similar to the chemiluminescence NO/
NO2 analyzers widely used to collect NO2
monitoring data for determining compliance with the NO2
NAAQS. The various oxides of nitrogen species, excluding NO, are first
quantitatively reduced to NO by means of a catalytic converter. These
species include NO2, HNO2, PANs, HNO3
and p-NO3. The NO, which commonly exists in ambient air,
passes through the converter unchanged, and, when combined with the NO
resulting from the catalytic conversion of the other oxides of
nitrogen, a measurement of the total NOy concentration
results. To maximize the conversion of the more chemically active
oxides of nitrogen species, the converter is located externally, at or
near the air sample inlet probe. This location minimizes losses of
these active species that could otherwise occur from chemical reactions
and wall losses in the sample inlet line.
The NOy analyzer is a suitable, commercially produced
continuous chemiluminescence analyzer that includes an ozone generator,
a reaction cell, a photometric detector, wavelength filters as
necessary to reduce sensitivity to wavelengths below 600 nanometer
(nm), a pump and flow control system to draw atmospheric air through
the converter and into the reaction cell, a suitable converter, a
system to control the operation of the analyzer, and appropriate
electronics to process and quantitatively scale the photometric
signals. The converter contains a catalyst such as molybdenum and is
heated to an optimum temperature designed to optimize the conversion of
the various oxides of nitrogen to NO. It is connected to the analyzer
via suitable lengths of Teflon[reg] tubing. Hourly NOy
measurements obtained by the analyzer would be averaged over the same
7-day period used by the CFP method to measure the SO2 and
p-SO4 components, with further averaging over a 1-year
period.
Commercial NOy analyzers are currently available, and
the analyzers have been used for a variety of monitoring applications.
During the 2006 TexAQS Radical and Aerosol Measurement Project (TRAMP),
Luke et al., 2010, compared measured NOy concentrations
obtained with an NOy instrument based upon the above
mentioned methodology with the sum of measured individual
NOy species (i.e., NOyi =
NO+NO2+HNO3+PANs+HNO2+p-
NO3). This comparison yielded excellent overall agreement
during both day ([NOy](ppb) = [NOyi](ppb) x 1.03-0.42; r\2\ = 0.9933)
and night time ([NOy](ppb) = [NOyi](ppb) x 1.01-0.18; r\2\ = 0.9975)
periods (Luke et al, 2010). The results of this study show that this
NOy method is capable of the accurate determination of all
the atmospherically relevant NOy components, resulting in an
accurate determination of total NOy concentrations. The
NOy instruments have been routinely operated in networks
such as SouthEastern Aerosol Research and Characterization (SEARCH),
dating back several years. In addition, state monitoring agencies
across the U.S. have begun, starting in 2009, the routine operation of
commercially available NOy instrumentation in anticipation
of EPA's NCore network transitioning to full operation in 2011.
These initial assessments described above are promising and
indicate that the photometric NOy method appears to be
accurate, reliable, and capable of routine network operation. As a
result, the method is most likely capable for use as an FRM for
determining atmospheric NOy concentrations as a component in
determining compliance with an AAI-based secondary standard. However,
as described below, this continuous method for NOy requires
additional time for further evaluation before it can be fully confirmed
for adoption as a FRM. The EPA has identified measurement uncertainties
and some remaining science questions associated with this method. Among
these are: (a) The ability of the method to capture all components of
NOy relevant to nitrogen deposition, (b) the efficiency of
the molybdenum converter in converting all oxides of nitrogen to NO for
detection (excluding NO2, as this conversion is already well
documented), (c) appropriate inlet height specifications to minimize
any bias associated with vertical concentration gradients of key
NOy components, (d) identification and quantification of
potential measurement interferences in the NOy
determination, and (e) development and demonstration of effective
calibration/challenge procedures to best represent the various mixtures
of NOy components that are expected to be present in the
different air sheds across the U.S.
To address these NOy method uncertainties and to fully
assess this method for use as the NOy FRM, EPA has developed
a detailed research plan (Russell and Samet, 2011b) which was presented
to the CASAC AMMS on February 16, 2011. In response, CASAC recognized
the need for, and supported the general outline of EPA's research plan
to evaluate the NOy method for potential designation as an
FRM (US EPA, 2011). In addition, the CASAC AMMS suggested additional
areas of research associated with the photometric NOy method
that warrant further assessment prior to final designation of the
method as the NOy FRM. These include operation of the method
during extremely low temperature conditions to investigate possible
condensation in sample lines, method detection limits relative to low
levels expected in remote areas, and ambient-based method evaluations
in various air sheds across the U.S. In response to these CASAC AMMS
suggestions, EPA is carrying out studies, in addition to the tasks
outlined in the research plan, for the NOy method. The
results of these studies will likely take a year or more to become
available. As noted previously, EPA anticipates that these results will
be favorable and will confirm the adequacy of the NOy method
as a suitable FRM for determining compliance with an AAI-based
secondary standard.

[[Page 46142]]

V. Statutory and Executive Order Reviews

A. Executive Order 12866: Regulatory Planning and Review and Review and
Executive Order 13563: Improving Regulation and Regulatory Review

Under Executive Order 12866 (58 FR 51735, October 4, 1993), this
action is a ``significant regulatory action.'' Accordingly, EPA
submitted this action to the Office of Management and Budget (OMB) for
review under Executive Orders 12866 and 13563 (76 FR 3821, January 21,
2011), and any changes made in response to OMB recommendations have
been documented in the docket for this action.

B. Paperwork Reduction Act

This action does not impose an information collection burden under
the provisions of the Paperwork Reduction Act, 44 U.S.C. 3501 et seq.
Burden is defined at 5 CFR 1320.3(b). There are no information
collection requirements directly associated with the establishment of a
NAAQS under section 109 of the CAA.

C. Regulatory Flexibility Act

For purposes of assessing the impacts of today's rule on small
entities, small entity is defined as: (1) A small business that is a
small industrial entity as defined by the Small Business
Administration's (SBA) regulations at 13 CFR 121.201; (2) a small
governmental jurisdiction that is a government of a city, county, town,
school district or special district with a population of less than
50,000; and (3) a small organization that is any not-for-profit
enterprise which is independently owned and operated and is not
dominant in its field.
After considering the economic impacts of today's proposed rule on
small entities, I certify that this action will not have a significant
economic impact on a substantial number of small entities. This
proposed rule will not impose any requirements on small entities.
Rather, this rule establishes national standards for allowable
concentrations of oxides of nitrogen and sulfur in ambient air as
required by section 109 of the CAA. See also American Trucking
Associations v. EPA. 175 F. 3d at 1044-45 (NAAQS do not have
significant impacts upon small entities because NAAQS themselves impose
no regulations upon small entities). We continue to be interested in
the potential impacts of the proposed rule on small entities and
welcome comments on issues related to such impacts.

D. Unfunded Mandates Reform Act

Title II of the Unfunded Mandates Reform Act of 1995 (UMRA), Public
Law 104-4, establishes requirements for Federal agencies to assess the
effects of their regulatory actions on State, local, and Tribal
governments and the private sector. Under section 202 of the UMRA, EPA
generally must prepare a written statement, including a cost-benefit
analysis, for proposed and final rules with ``Federal mandates'' that
may result in expenditures to state, local, and tribal governments, in
the aggregate, or to the private sector, of $100 million or more in any
1 year. Before promulgating an EPA rule for which a written statement
is needed, section 205 of the UMRA generally requires EPA to identify
and consider a reasonable number of regulatory alternatives and to
adopt the least costly, most cost-effective or least burdensome
alternative that achieves the objectives of the rule. The provisions of
section 205 do not apply when they are inconsistent with applicable
law. Moreover, section 205 allows EPA to adopt an alternative other
than the least costly, most cost-effective or least burdensome
alternative if the Administrator publishes with the final rule an
explanation why that alternative was not adopted. Before EPA
establishes any regulatory requirements that may significantly or
uniquely affect small governments, including tribal governments, it
must have developed under section 203 of the UMRA a small government
agency plan. The plan must provide for notifying potentially affected
small governments, enabling officials of affected small governments to
have meaningful and timely input in the development of EPA regulatory
proposals with significant Federal intergovernmental mandates, and
informing, educating, and advising small governments on compliance with
the regulatory requirements.
This action contains no Federal mandates under the provisions of
Title II of the Unfunded Mandates Reform Act of 1995 (UMRA), 2 U.S.C.
1531-1538 for state, local, or tribal governments or the private
sector. Therefore, this action is not subject to the requirements of
sections 202 or 205. Furthermore, as indicated previously, in setting a
NAAQS EPA cannot consider the economic or technological feasibility of
attaining ambient air quality standards; although such factors may be
considered to a degree in the development of state plans to implement
the standards. See also American Trucking Associations v. EPA, 175 F.
3d at 1043 (noting that because EPA is precluded from considering costs
of implementation in establishing NAAQS, preparation of a Regulatory
Impact Analysis pursuant to the Unfunded Mandates Reform Act would not
furnish any information which the court could consider in reviewing the
NAAQS). Accordingly, EPA has determined that the provisions of sections
202, 203, and 205 of the UMRA do not apply to this proposed decision.
The EPA acknowledges, however, that any corresponding revisions to
associated state implementation plan (SIP) requirements and air quality
surveillance requirements, 40 CFR part 51 and 40 CFR part 58,
respectively, might result in such effects. Accordingly, EPA will
address, as appropriate, unfunded mandates if and when it proposes any
revisions to 40 CFR parts 51 or 58.

E. Executive Order 13132: Federalism

This proposed rule does not have federalism implications. It will
not have substantial direct effects on the states, on the relationship
between the national government and the states, or on the distribution
of power and responsibilities among the various levels of government,
as specified in Executive Order 13132 because it does not contain
legally binding requirements. Thus, the requirements of Executive Order
13132 do not apply to this rule.
EPA believes, however, that this proposed rule may be of
significant interest to state governments. As also noted in section E
(above) on UMRA, EPA recognizes that states will have a substantial
interest in this rule and any corresponding revisions to associated SIP
requirements and air quality surveillance requirements, 40 CFR part 51
and 40 CFR part 58, respectively. Therefore, in the spirit of Executive
Order 13132 and consistent with EPA policy to promote communications
between EPA and state and local governments, EPA specifically solicits
comment on this proposed rule from state and local officials.

F. Executive Order 13175: Consultation and Coordination With Indian
Tribal Governments

Executive Order 13175, entitled ``Consultation and Coordination
with Indian Tribal Governments'' (65 FR 67249, November 9, 2000),
requires EPA to develop an accountable process to ensure ``meaningful
and timely input by tribal officials in the development of regulatory
policies that have tribal implications.'' This rule concerns the
establishment of national standards to address the public welfare
effects of oxides of nitrogen and sulfur.

[[Page 46143]]

This action does not have Tribal implications, as specified in
Executive Order 13175 (65 FR 67249, November 9, 2000). It does not have
a substantial direct effect on one or more Indian tribes, since tribes
are not obligated to adopt or implement any NAAQS. Thus, Executive
Order 13175 does not apply to this rule.

G. Executive Order 13045: Protection of Children from Environmental
Health & Safety Risks

This action is not subject to EO 13045 because it is not an
economically significant rule as defined in EO 12866.

H. Executive Order 13211: Actions that Significantly Affect Energy
Supply, Distribution or Use

This action is not a ``significant energy action'' as defined in
Executive Order 13211 (66 FR 28355, May 22, 2001), because it is not
likely to have a significant adverse effect on the supply,
distribution, or use of energy. This action concerns the establishment
of national standards to address the public welfare effects of oxides
of nitrogen and sulfur. This action does not prescribe specific
pollution control strategies by which these ambient standards will be
met. Such strategies will be developed by states on a case-by-case
basis, and EPA cannot predict whether the control options selected by
states will include regulations on energy suppliers, distributors, or
users.

I. National Technology Transfer and Advancement Act

Section 12(d) of the National Technology Transfer and Advancement
Act of 1995 (NTTAA), Public Law 104-113, 12(d) (15 U.S.C. 272 note)
directs EPA to use voluntary consensus standards in its regulatory
activities unless to do so would be inconsistent with applicable law or
otherwise impractical. Voluntary consensus standards are technical
standards (e.g., materials specifications, test methods, sampling
procedures, and business practices) that are developed or adopted by
voluntary consensus standards bodies. The NTTAA directs EPA to provide
Congress, through OMB, explanations when the Agency decides not to use
available and applicable voluntary consensus standards.
The EPA is not aware of any voluntary consensus standards that are
relevant to the provisions of this proposed rule. The EPA welcomes any
feedback on such standards that may be applicable.

J. Executive Order 12898: Federal Actions To Address Environmental
Justice in Minority Populations and Low-Income Populations

Executive Order 12898 (59 FR 7629 (Feb. 16, 1994)) establishes
federal executive policy on environmental justice. Its main provision
directs federal agencies, to the greatest extent practicable and
permitted by law, to make environmental justice part of their mission
by identifying and addressing, as appropriate, disproportionately high
and adverse human health or environmental effects of their programs,
policies, and activities on minority populations and low-income
populations in the United States.
EPA has determined that this proposed rule will not have
disproportionately high and adverse human health or environmental
effects on minority or low-income populations because it retains the
level of environmental protection for all affected populations without
having any disproportionately high and adverse human health or
environmental effects on any population, including any minority or low-
income population.

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effects of changes in surface water acid-base chemistry. (State of
science/technology report 13). Washington DC: National Acid
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and projected future status of brook trout (Salvelinus fontinalis)
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Banzhaf, S., D. Burtraw, D. Evans, and A. Krupnick. 2006.
``Valuation of Natural Resource Improvements in the Adirondacks.''
Land Economics 82:445-464.
Lavery, T.F. C.M. Rogers, R. Baumgardner, and K.P. Mishoe. 2009.
Intercomparison of Clean Air Status and Trends Network Nitrate and
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waters: invertebrates and fish. (Acid rain research report no 21).
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National Acid Precipitation Assessment Program. Office of the
Director, Washington, DC.
NAPAP. (2005). National acid precipitation assessment program
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Committee on Environment and Natural Resources (CENR) of the
National Science and Technology Council (NSTC).
NRC (National Research Council). 2004. Air quality management in
the United States. Washington, DC: National Research Council (NRC);
The National Academies Press.
Russell, A and J. M. Samet, 2010a. Review of the Policy
Assessment for the Review of the Secondary National Ambient Air
Quality Standard for NOX and SOX: First Draft.
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Russell, A and J. M. Samet, 2010b. Review of the Policy
Assessment for the Review of the Secondary National Ambient Air
Quality Standard for NOX and SOX: Second
Draft. EPA-CASAC-11-003.
Russell, A and J. M. Samet, 2011. Review of the Policy
Assessment for the Review of the Secondary National Ambient Air
Quality Standard for NOX and SOX: FINAL. EPA-
CASAC-11-005.
Russell and Samet, 2011b Review of EPA Draft Documents on
Monitoring and Methods for Oxides of Nitrogen (NOX) and
Sulfur (SOX) http://yosemite.epa.gov/sab/sabpeople.nsf/WebCommittees/CASAC.
Sickles II, J.E., L. L. Hodson, and L. M. Vorburger. 1999.
Evaluation of the filter pack for long-duration sampling of ambient
air. Atmospheric Environment, 33, 2187-2202.
Sickles II, J.E. and D.S. Shadwick. 2008. Comparison of
particulate sulfate and nitrate at collocated CASTNET and IMPROVE
sites in the eastern US. Atmospheric Environment, 42, 2062-2073.
Smyth SC, W. Jiang, and H. Roth. 2008. A comparative performance
evaluation of the AURAMS and CMAQ air quality modeling systems.
Atmos Envir 43:1059-1070.
Stoddard J; Kahl JS; Deviney FA; DeWalle DR; Driscoll CT;
Herlihy AT; Kellogg JH; Murdoch PS; Webb JR; Webster KE. (2003).
Response of surface water chemistry to the Clean Air Act Amendments
of 1990 (No. EPA 620/R-03/001). Research Triangle Park, NC; National
Health and Environmental Effects Research Laboratory; Office of
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Sullivan TJ; Driscoll CT; Cosby BJ; Fernandez IJ; Herlihy AT;
Zhai J; Stemberger R; Snyder KU; Sutherland JW; Nierzwicki-Bauer SA;
Boylen CW; McDonnell TC; Nowicki NA. 2006. Assessment of the extent
to which intensively studied lakes are representative of the
Adirondack Mountain region. (Final Report no 06-17).Corvallis, OR;
prepared by Environmental Chemistry, Inc. for: Albany, NY;
Environmental Monitoring Evaluation and Protection Program of the
New York State Energy Research and Development Authority (NYSERDA).

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US EPA, 1973. ``Effects of Sulfur Oxide in the Atmosphere on
Vegetation''. Revised Chapter 5 of Air Quality Criteria For Sulfur
Oxides. U.S. Environmental Protection Agency. Research Triangle
Park, N.C. EPA-R3-73-030.
US EPA. 1982. Review of the National Ambient Air Quality
Standards for Sulfur Oxides: Assessment of Scientific and Technical
Information. OAQPS Staff Paper. EPA-450/5-82-007. U.S. Environmental
Protection Agency, Office of Air Quality Planning and Standards,
Research Triangle Park, NC.
US EPA, 1984a. The Acidic Deposition Phenomenon and Its Effects:
Critical Assessment Review Papers. Volume I Atmospheric Sciences.
EPA-600/8-83-016AF. Office of Research and Development, Washington,
DC.
US EPA, 1984b. The Acidic Deposition Phenomenon and Its Effects:
Critical Assessment Review Papers. Volume II Effects Sciences. EPA-
600/8-83-016BF. Office of Research and Development, Washington, DC.
US EPA, 1985. The Acidic Deposition Phenomenon and Its Effects:
Critical Assessment Document. EPA-600/8-85/001. Office of Research
and Development, Washington, DC.
US EPA. 1995a. Review of the National Ambient Air Quality
Standards for Nitrogen Dioxide: Assessment of Scientific and
Technical Information. OAQPS Staff Paper. EPA-452/R-95-005. U.S.
Environmental Protection Agency, Office of Air Quality Planning and
Standards, Research Triangle Park, NC. September.
US EPA. 1995b. Acid Deposition Standard Feasibility Study Report
to Congress. U.S. Environmental Protection Agency, Washington, DC.
EPA-430/R-95-001a.
US EPA 2007. Integrated Review Plan for the Secondary National
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Dioxide. U.S. Environmental Protection Agency, Research Triangle
Park, NC, EPA-452/R-08-006.
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Nitrogen and Sulfur Ecological Criteria (Final Report). U.S.
Environmental Protection Agency, Washington, D.C., EPA/600/R-08/
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Secondary National Ambient Air Quality Standards for Oxides of
Nitrogen and Oxides of Sulfur-Main Content--Final Report. U.S.
Environmental Protection Agency, Washington, D.C., EPA-452/R-09-
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7.0, October 2010, http://java.epa.gov/castnet/.
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US EPA 2011. Policy Assessment for the Review of the Secondary
National Ambient Air Quality Standards for Oxides of Nitrogen and
Oxides of Sulfur. U.S. Environmental Protection Agency, Washington,
DC, EPA-452/R-11-005a.
US EPA, 2011b. Federal Reference Methods for NOy and
p-SO4 for the New Combined NOX and SOx
Secondary NAAQS Research Plan, EPA/600/1-11/002 January 20, 2011.
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August 22, 1995.

List of Subjects in 40 CFR Part 50

Environmental protection, Air pollution control, Carbon monoxide,
Lead, Nitrogen dioxide, Ozone, Particulate matter, Sulfur oxides.

Dated: July 12, 2011.
Lisa P. Jackson,
Administrator.

For the reasons set forth in the preamble, part 50 of chapter 1 of
title 40 of the code of Federal regulations is proposed to be amended
as follows:

PART 50--NATIONAL PRIMARY AND SECONDARY AMBIENT AIR QUALITY
STANDARDS

1. The authority citation for part 50 continues to read as follows:

Authority: 42 U.S.C. 7401, et seq.
2. Section 50.5 is amended by revising paragraphs (b) and (c) and
by adding paragraphs (d) and (e) to read as follows:

Sec. 50.5 National secondary ambient air quality standards for sulfur
oxides (sulfur dioxide).

* * * * *
(b) The level of the national secondary 1-hour ambient air quality
standard for oxides of sulfur is 75 parts per billion (ppb, which is 1
part in 1,000,000,000), measured in the ambient air as sulfur dioxide
(SO2).
(c) The levels of the standards shall be measured by a reference
method based on Appendix A-1 or A-2 of this part, or by a Federal
Equivalent Method (FEM) designated in accordance with part 53 of this
chapter.
(d) To demonstrate attainment with the 3-hour secondary standard,
the second-highest 3-hour average must be based upon hourly data that
are at least 75 percent complete in each calendar quarter. A 3-hour
block average shall be considered valid only if all three hourly
averages for the 3-hour period are available. If only one or two hourly
averages are available, but the 3-hour average would exceed the level
of the standard when zeros are substituted for the missing values,
subject to the rounding rule of paragraph (a) of this section, then
this shall be considered a valid 3-hour average. In all cases, the 3-
hour block average shall be computed as the sum of the hourly averages
divided by 3.
(e) The 1-hour secondary standard is met at an ambient air quality
monitoring site when the three-year average of the annual 99th
percentile of the daily maximum 1-hour average concentrations is less
than or equal to 75 ppb, as determined in accordance with Appendix T of
this part.
3. Section 50.11 is revised to read as follows:

Sec. 50.11 National primary and secondary ambient air quality
standards for oxides of nitrogen (with nitrogen dioxide as the
indicator).

(a) The level of the national primary and secondary annual ambient
air quality standards for oxides of nitrogen is 53 parts per billion
(ppb, which is 1 part in 1,000,000,000), annual average concentration,
measured in the ambient air as nitrogen dioxide.
(b) The level of the national primary and secondary 1-hour ambient
air quality standards for oxides of nitrogen is 100 ppb, 1-hour average
concentration, measured in the ambient air as nitrogen dioxide.
(c) The levels of the standards shall be measured by:
(1) A reference method based on appendix F to this part; or
(2) A Federal equivalent method (FEM) designated in accordance with
part 53 of this chapter.
(d) The annual primary and secondary standards are met when the
annual average concentration in a calendar year is less than or equal
to 53 ppb, as determined in accordance with Appendix S of this part for
the annual standard.
(e) The 1-hour primary and secondary standards are met when the
three-year average of the annual 98th percentile of the daily maximum
1-hour average concentration is less than or equal to 100 ppb, as
determined in accordance with Appendix S of this part for the 1-hour
standard.
4. Appendix S is amended as follows:
a. by revising paragraph 1.(a),
b. by revising the definition of ``Design values'' under paragraph
1.(c),
c. by revising paragraph 2.(b),
d. by revising paragraphs 3.1(a) through (d),
e. by revising paragraphs 3.2(a) through (e),
f. by revising paragraph 4.1(b),
g. by revising paragraph 4.2(c),
h. by revising paragraph 5.1(b), and
i. by revising paragraph 5.2(b) to read as follows:

[[Page 46145]]

Appendix S to Part 50--Interpretation of the Primary and Secondary
National Ambient Air Quality Standards for Oxides of Nitrogen (Nitrogen
Dioxide)

1. General.
(a) This appendix explains the data handling conventions and
computations necessary for determining when the primary and
secondary national ambient air quality standards for oxides of
nitrogen as measured by nitrogen dioxide (``NO2 NAAQS'')
specified in Sec. 50.11 are met. Nitrogen dioxide (NO2)
is measured in the ambient air by a Federal reference method (FRM)
based on appendix F to this part or by a Federal equivalent method
(FEM) designated in accordance with part 53 of this chapter. Data
handling and computation procedures to be used in making comparisons
between reported NO2 concentrations and the levels of the
NO2 NAAQS are specified in the following sections.
* * * * *
(c) * * *
Design values are the metrics (i.e., statistics) that are
compared to the NAAQS levels to determine compliance, calculated as
specified in section 5 of this appendix. The design values for the
primary and secondary NAAQS are:
(1) The annual mean value for a monitoring site for one year
(referred to as the ``annual primary or secondary standard design
value'').
(2) The 3-year average of annual 98th percentile daily maximum
1-hour values for a monitoring site (referred to as the ``1-hour
primary or secondary standard design value'').
* * * * *
2. Requirements for Data Used for Comparisons With the
NO2 NAAQS and Data
Reporting Considerations.
* * * * *
(b) When two or more NO2 monitors are operated at a
site, the state may in advance designate one of them as the primary
monitor. If the state has not made this designation, the
Administrator will make the designation, either in advance or
retrospectively. Design values will be developed using only the data
from the primary monitor, if this results in a valid design value.
If data from the primary monitor do not allow the development of a
valid design value, data solely from the other monitor(s) will be
used in turn to develop a valid design value, if this results in a
valid design value. If there are three or more monitors, the order
for such comparison of the other monitors will be determined by the
Administrator. The Administrator may combine data from different
monitors in different years for the purpose of developing a valid 1-
hour primary or secondary standard design value, if a valid design
value cannot be developed solely with the data from a single
monitor. However, data from two or more monitors in the same year at
the same site will not be combined in an attempt to meet data
completeness requirements, except if one monitor has physically
replaced another instrument permanently, in which case the two
instruments will be considered to be the same monitor, or if the
state has switched the designation of the primary monitor from one
instrument to another during the year.
* * * * *
3. Comparisons with the NO2 NAAQS.
3.1 The Annual Primary and Secondary NO2 NAAQS.
(a) The annual primary and secondary NO2 NAAQS are
met at a site when the valid annual primary standard design value is
less than or equal to 53 parts per billion (ppb).
(b) An annual primary or secondary standard design value is
valid when at least 75 percent of the hours in the year are
reported.
(c) An annual primary or secondary standard design value based
on data that do not meet the completeness criteria stated in section
3.1(b) may also be considered valid with the approval of, or at the
initiative of, the Administrator, who may consider factors such as
monitoring site closures/moves, monitoring diligence, the
consistency and levels of the valid concentration measurements that
are available, and nearby concentrations in determining whether to
use such data.
(d) The procedures for calculating the annual primary and
secondary standard design values are given in section 5.1 of this
appendix.
3.2 The 1-Hour Primary and Secondary NO2 NAAQS.
(a) The 1-hour primary or secondary NO2 NAAQS is met
at a site when the valid 1-hour primary or secondary standard design
value is less than or equal to 100 parts per billion (ppb).
(b) An NO2 1-hour primary or secondary standard
design value is valid if it encompasses three consecutive calendar
years of complete data. A year meets data completeness requirements
when all 4 quarters are complete. A quarter is complete when at
least 75 percent of the sampling days for each quarter have complete
data. A sampling day has complete data if 75 percent of the hourly
concentration values, including state-flagged data affected by
exceptional events which have been approved for exclusion by the
Administrator, are reported.
(c) In the case of one, two, or three years that do not meet the
completeness requirements of section 3.2(b) of this appendix and
thus would normally not be useable for the calculation of a valid 3-
year 1-hour primary or secondary standard design value, the 3-year
1-hour primary or secondary standard design value shall nevertheless
be considered valid if one of the following conditions is true.
(i) At least 75 percent of the days in each quarter of each of
three consecutive years have at least one reported hourly value, and
the design value calculated according to the procedures specified in
section 5.2 is above the level of the primary or secondary 1-hour
standard.
(ii) (A) A 1-hour primary or secondary standard design value
that is below the level of the NAAQS can be validated if the
substitution test in section 3.2(c)(ii)(B) results in a ``test
design value'' that is below the level of the NAAQS. The test
substitutes actual ``high'' reported daily maximum 1-hour values
from the same site at about the same time of the year (specifically,
in the same calendar quarter) for unknown values that were not
successfully measured. Note that the test is merely diagnostic in
nature, intended to confirm that there is a very high likelihood
that the original design value (the one with less than 75 percent
data capture of hours by day and of days by quarter) reflects the
true under-NAAQS-level status for that 3-year period; the result of
this data substitution test (the ``test design value'', as defined
in section 3.2(c)(ii)(B)) is not considered the actual design value.
For this test, substitution is permitted only if there are at least
200 days across the three matching quarters of the three years under
consideration (which is about 75 percent of all possible daily
values in those three quarters) for which 75 percent of the hours in
the day, including state-flagged data affected by exceptional events
which have been approved for exclusion by the Administrator, have
reported concentrations. However, maximum 1-hour values from days
with less than 75 percent of the hours reported shall also be
considered in identifying the high value to be used for
substitution.
(B) The substitution test is as follows: Data substitution will
be performed in all quarter periods that have less than 75 percent
data capture but at least 50 percent data capture, including state-
flagged data affected by exceptional events which have been approved
for exclusion by the Administrator; if any quarter has less than 50
percent data capture then this substitution test cannot be used.
Identify for each quarter (e.g., January-March) the highest reported
daily maximum 1-hour value for that quarter, excluding state-flagged
data affected by exceptional events which have been approved for
exclusion by the Administrator, looking across those three months of
all three years under consideration. All daily maximum 1-hour values
from all days in the quarter period shall be considered when
identifying this highest value, including days with less than 75
percent data capture. If after substituting the highest non-excluded
reported daily maximum 1-hour value for a quarter for as much of the
missing daily data in the matching deficient quarter(s) as is needed
to make them 100 percent complete, the procedure in section 5.2
yields a recalculated 3-year 1-hour standard ``test design value''
below the level of the standard, then the 1-hour primary or
secondary standard design value is deemed to have passed the
diagnostic test and is valid, and the level of the standard is
deemed to have been met in that 3-year period. As noted in section
3.2(c)(i), in such a case, the 3-year design value based on the data
actually reported, not the ``test design value'', shall be used as
the valid design value. (iii) (A) A 1-hour primary or secondary
standard design value that is above the level of the NAAQS can be
validated if the substitution test in section 3.2(c)(iii)(B) results
in a ``test design value'' that is above the level of the NAAQS. The
test substitutes actual ``low'' reported daily maximum 1-hour values
from the same site at about the same time of the year (specifically,
in the same three months of the

[[Page 46146]]

calendar) for unknown values that were not successfully measured.
Note that the test is merely diagnostic in nature, intended to
confirm that there is a very high likelihood that the original
design value (the one with less than 75 percent data capture of
hours by day and of days by quarter) reflects the true above-NAAQS-
level status for that 3-year period; the result of this data
substitution test (the ``test design value'', as defined in section
3.2(c)(iii)(B)) is not considered the actual design value. For this
test, substitution is permitted only if there are a minimum number
of available daily data points from which to identify the low
quarter-specific daily maximum 1-hour values, specifically if there
are at least 200 days across the three matching quarters of the
three years under consideration (which is about 75 percent of all
possible daily values in those three quarters) for which 75 percent
of the hours in the day have reported concentrations. Only days with
at least 75 percent of the hours reported shall be considered in
identifying the low value to be used for substitution.
(B) The substitution test is as follows: Data substitution will
be performed in all quarter periods that have less than 75 percent
data capture. Identify for each quarter (e.g., January-March) the
lowest reported daily maximum 1-hour value for that quarter, looking
across those three months of all three years under consideration.
All daily maximum 1-hour values from all days with at least 75
percent capture in the quarter period shall be considered when
identifying this lowest value. If after substituting the lowest
reported daily maximum 1-hour value for a quarter for as much of the
missing daily data in the matching deficient quarter(s) as is needed
to make them 75 percent complete, the procedure in section 5.2
yields a recalculated 3-year 1-hour standard ``test design value''
above the level of the standard, then the 1-hour primary or
secondary standard design value is deemed to have passed the
diagnostic test and is valid, and the level of the standard is
deemed to have been exceeded in that 3-year period. As noted in
section 3.2(c)(i), in such a case, the 3-year design value based on
the data actually reported, not the ``test design value'', shall be
used as the valid design value.
(d) A 1-hour primary or secondary standard design value based on
data that do not meet the completeness criteria stated in 3.2(b) and
also do not satisfy section 3.2(c), may also be considered valid
with the approval of, or at the initiative of, the Administrator,
who may consider factors such as monitoring site closures/moves,
monitoring diligence, the consistency and levels of the valid
concentration measurements that are available, and nearby
concentrations in determining whether to use such data.
(e) The procedures for calculating the 1-hour primary and
secondary standard design values are given in section 5.2 of this
appendix.
4. Rounding Conventions.
4.1 Rounding Conventions for the Annual Primary and Secondary
NO2 NAAQS.
* * * * *
(b) The annual primary or secondary standard design value is
calculated pursuant to section 5.1 and then rounded to the nearest
whole number or 1 ppb (decimals 0.5 and greater are rounded up to
the nearest whole number, and any decimal lower than 0.5 is rounded
down to the nearest whole number).
4.2 Rounding Conventions for the 1-hour Primary and Secondary
NO2 NAAQS.
* * * * *
(c) The 1-hour primary or secondary standard design value is
calculated pursuant to section 5.2 and then rounded to the nearest
whole number or 1 ppb (decimals 0.5 and greater are rounded up to
the nearest whole number, and any decimal lower than 0.5 is rounded
down to the nearest whole number).
5. Calculation Procedures for the Primary and Secondary
NO2 NAAQS.
5.1 Procedures for the Annual Primary and Secondary
NO2 NAAQS.
* * * * *
(b) The annual primary or secondary standard design value for a
site is the valid annual mean rounded according to the conventions
in section 4.1.
5.2 Calculation Procedures for the 1-hour Primary and Secondary
NO2 NAAQS.
* * * * *
(b) The 1-hour primary or secondary standard design value for a
site is the mean of the three annual 98th percentile values, rounded
according to the conventions in section 4.
* * * * *

5. Appendix T is amended as follows:
a. by revising paragraph 1.(a),
b. by revising the definition of ``Design values'' under paragraph
1.(c),
c. by revising paragraph 2.(b),
d. by revising paragraphs 3.(a) through (e),
e. by revising paragraph 4.(c), and
f. by revising paragraph 5.(b) to read as follows:

Appendix T to Part 50--Interpretation of the Primary and Secondary
National Ambient Air Quality Standards for Oxides of Sulfur (Sulfur
Dioxide)

[[Page 46147]]

(c) In the case of one, two, or three years that do not meet the
completeness requirements of section 3(b) of this appendix and thus
would normally not be useable for the calculation of a valid 3-year
1-hour primary or secondary standard design value, the 3-year 1-hour
primary or secondary standard design value shall nevertheless be
considered valid if one of the following conditions is true.
(i) At least 75 percent of the days in each quarter of each of
three consecutive years have at least one reported hourly value, and
the design value calculated according to the procedures specified in
section 5 is above the level of the primary or secondary 1-hour
standard.
(ii) (A) A 1-hour primary or secondary standard design value
that is equal to or below the level of the NAAQS can be validated if
the substitution test in section 3(c)(ii)(B) results in a ``test
design value'' that is below the level of the NAAQS. The test
substitutes actual ``high'' reported daily maximum 1-hour values
from the same site at about the same time of the year (specifically,
in the same calendar quarter) for unknown values that were not
successfully measured. Note that the test is merely diagnostic in
nature, intended to confirm that there is a very high likelihood
that the original design value (the one with less than 75 percent
data capture of hours by day and of days by quarter) reflects the
true under-NAAQS-level status for that 3-year period; the result of
this data substitution test (the ``test design value'', as defined
in section 3(c)(ii)(B)) is not considered the actual design value.
For this test, substitution is permitted only if there are at least
200 days across the three matching quarters of the three years under
consideration (which is about 75 percent of all possible daily
values in those three quarters) for which 75 percent of the hours in
the day, including State-flagged data affected by exceptional events
which have been approved for exclusion by the Administrator, have
reported concentrations. However, maximum 1-hour values from days
with less than 75 percent of the hours reported shall also be
considered in identifying the high value to be used for
substitution.
(B) The substitution test is as follows: Data substitution will
be performed in all quarter periods that have less than 75 percent
data capture but at least 50 percent data capture, including State-
flagged data affected by exceptional events which have been approved
for exclusion by the Administrator; if any quarter has less than 50
percent data capture then this substitution test cannot be used.
Identify for each quarter (e.g., January-March) the highest reported
daily maximum 1-hour value for that quarter, excluding State-flagged
data affected by exceptional events which have been approved for
exclusion by the Administrator, looking across those three months of
all three years under consideration. All daily maximum 1-hour values
from all days in the quarter period shall be considered when
identifying this highest value, including days with less than 75
percent data capture. If after substituting the highest reported
daily maximum 1-hour value for a quarter for as much of the missing
daily data in the matching deficient quarter(s) as is needed to make
them 100 percent complete, the procedure in section 5 yields a
recalculated 3-year 1-hour standard ``test design value'' less than
or equal to the level of the standard, then the 1-hour primary or
secondary standard design value is deemed to have passed the
diagnostic test and is valid, and the level of the standard is
deemed to have been met in that 3-year period. As noted in section
3(c)(i), in such a case, the 3-year design value based on the data
actually reported, not the ``test design value'', shall be used as
the valid design value.
(iii) (A) A 1-hour primary or secondary standard design value
that is above the level of the NAAQS can be validated if the
substitution test in section 3(c)(iii)(B) results in a ``test design
value'' that is above the level of the NAAQS. The test substitutes
actual ``low'' reported daily maximum 1-hour values from the same
site at about the same time of the year (specifically, in the same
three months of the calendar) for unknown hourly values that were
not successfully measured. Note that the test is merely diagnostic
in nature, intended to confirm that there is a very high likelihood
that the original design value (the one with less than 75 percent
data capture of hours by day and of days by quarter) reflects the
true above-NAAQS-level status for that 3-year period; the result of
this data substitution test (the ``test design value'', as defined
in section 3(c)(iii)(B)) is not considered the actual design value.
For this test, substitution is permitted only if there are a minimum
number of available daily data points from which to identify the low
quarter-specific daily maximum 1-hour values, specifically if there
are at least 200 days across the three matching quarters of the
three years under consideration (which is about 75 percent of all
possible daily values in those three quarters) for which 75 percent
of the hours in the day have reported concentrations. Only days with
at least 75 percent of the hours reported shall be considered in
identifying the low value to be used for substitution.
(B) The substitution test is as follows: Data substitution will
be performed in all quarter periods that have less than 75 percent
data capture. Identify for each quarter (e.g., January-March) the
lowest reported daily maximum 1-hour value for that quarter, looking
across those three months of all three years under consideration.
All daily maximum 1-hour values from all days with at least 75
percent capture in the quarter period shall be considered when
identifying this lowest value. If after substituting the lowest
reported daily maximum 1-hour value for a quarter for as much of the
missing daily data in the matching deficient quarter(s) as is needed
to make them 75 percent complete, the procedure in section 5 yields
a recalculated 3-year 1-hour standard ``test design value'' above
the level of the standard, then the 1-hour primary or secondary
standard design value is deemed to have passed the diagnostic test
and is valid, and the level of the standard is deemed to have been
exceeded in that 3-year period. As noted in section 3(c)(i), in such
a case, the 3-year design value based on the data actually reported,
not the ``test design value'', shall be used as the valid design
value.
(d) A 1-hour primary or secondary standard design value based on
data that do not meet the completeness criteria stated in 3(b) and
also do not satisfy section 3(c), may also be considered valid with
the approval of, or at the initiative of, the Administrator, who may
consider factors such as monitoring site closures/moves, monitoring
diligence, the consistency and levels of the valid concentration
measurements that are available, and nearby concentrations in
determining whether to use such data.
(e) The procedures for calculating the 1-hour primary or
secondary standard design values are given in section 5 of this
appendix.
4. Rounding Conventions for the 1-hour Primary and Secondary
SO2 NAAQS.
* * * * *
(c) The 1-hour primary or secondary standard design value is
calculated pursuant to section 5 and then rounded to the nearest
whole number or 1 ppb (decimals 0.5 and greater are rounded up to
the nearest whole number, and any decimal lower than 0.5 is rounded
down to the nearest whole number).
5. Calculation Procedures for the 1-hour Primary and Secondary
SO2 NAAQS.
* * * * *
(b) The 1-hour primary or secondary standard design value for an
ambient air quality monitoring site is the mean of the three annual
99th percentile values, rounded according to the conventions in
section 4.

[FR Doc. 2011-18582 Filed 7-29-11; 8:45 am]
BILLING CODE 6560-50-P