ENVIRONMENTAL PROTECTION AGENCY

40 CFR Part 50

[EPA-HQ-OAR-2007-1145]

RIN: 2060-AO72

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

AGENCY:  Environmental Protection Agency (EPA).

ACTION:  Proposed rule.

Instead 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 standard 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.

DATES:  Written comments on this proposed rule must be received by
[insert date 60 days after date of publication in the Federal Register].

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:

 HYPERLINK "http://www.regulations.gov" www.regulations.gov :  Follow
the on-line instructions for submitting comments.

Email:   HYPERLINK "mailto:a-and-r-Docket@epa.gov"
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  HYPERLINK "http://www.regulations.gov"
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  HYPERLINK
"http://www.regulations.gov" www.regulations.gov  or email.  The 
HYPERLINK "http://www.regulations.gov" www.regulations.gov  website 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 email comment directly to EPA without
going through  HYPERLINK "http://www.regulations.gov"
www.regulations.gov , your email 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  HYPERLINK
"http://www.epa.gov/epahome/dockets.htm"
http://www.epa.gov/epahome/dockets.htm . 

	Docket:  All documents in the docket are listed in the  HYPERLINK
"http://www.regulations.gov" 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  HYPERLINK
"http://www.regulations.gov" 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; email: scheffe.rich@epa.gov. 

SUPPLEMENTARY INFORMATION:

General Information

What Should I Consider as I Prepare My Comments for EPA?

Submitting CBI.  Do not submit this information to EPA through 
HYPERLINK "http://www.regulations.gov" www.regulations.gov  or email. 
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.

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).

 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.

.

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  HYPERLINK "http://www.epa.gov/ncea"
http://www.epa.gov/ncea , and click on Air Quality then click on Oxides
of Nitrogen and Sulfur.  The Policy Assessment, 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 Policy Assessment 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  HYPERLINK
"http://www.epa.gov/ttn/naaqs/standards/no2so2sec/cr_rea.html"
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 

Ecological Effects

Effects Associated with Gas-Phase Oxides of Nitrogen and Sulfur

Nature of ecosystem responses to gas-phase nitrogen and sulfur

Magnitude of ecosystem response to gas-phase nitrogen and sulfur

Acidification Effects Associated with Deposition of Oxides of Nitrogen 	
	and Sulfur

Nature of Acidification –related Ecosystem Responses

Aquatic ecosystems

Terrestrial ecosystems

Ecosystem sensitivity

Magnitude of Acidification-related Ecosystem Responses

Aquatic acidification

Terrestrial acidification

Key Uncertainties Associated with Acidification

Aquatic acidification

Terrestrial acidification

Nutrient Enrichment Effects Associated with Deposition of Oxides of 
Nitrogen 

Nature of Nutrient Enrichment-related Ecosystem Responses

	i.	Aquatic ecosystems

	ii.	Terrestrial ecosystems

	iii.	Ecosystem sensitivity to nutrient enrichment

Magnitude of Nutrient Enrichment-related Ecosystem Responses

	i.	Aquatic ecosystems

	ii.	Terrestrial ecosystems

Key Uncertainties Associated with Nutrient Enrichment

Aquatic ecosystems

Terrestrial ecosystems

4.	Other Ecological Effects

Risk and Exposure Assessment

Overview of Risk and Exposure Assessment

Key Findings

Air quality analyses

Deposition-related aquatic acidification

Deposition-related terrestrial acidification

Deposition-related aquatic nutrient enrichment

Deposition-related terrestrial nutrient enrichment 

Additional effects

Conclusions on Effects

Adversity of Effects to Public Welfare

Ecosystem Services

Effects on Ecosystem Services

Aquatic Acidification

Terrestrial Acidification

Nutrient Enrichment

Summary

Adequacy of the Current Standards

Adequacy of the Current Standards for Direct Effects

Appropriateness and Adequacy of the Current Standards for
Deposition-related Effects

Appropriateness

Adequacy of Protection

Aquatic Acidification

Terrestrial acidification

Terrestrial nutrient enrichment

Aquatic nutrient enrichment

Other effects

CASAC Views

Administrator’s Proposed Conclusions Concerning Adequacy of Current
Standard

Rationale for Proposed Decision on Alternative Multi-pollutant Approach
to Secondary Standards for Aquatic Acidification

Ambient Air Indicators

Oxides of Sulfur

Oxides of Nitrogen

Form

Ecological Indicator

Linking ANC to Deposition

Linking Deposition to Ambient Air Indicators

Aquatic Acidification Index

Spatial Aggregation

Ecoregion Sensitivity

Representative Ecoregion-specific Factors

Factor F1

Acid-sensitive Ecoregions

Non-acid sensitive Ecoregions

Factor F2

Factors F3 and F4

Factors in Data-limited Ecoregions

Application to Hawaii, Alaska, and the U.S. Territories

Summary of the AAI Form

Averaging Time

Level

Association Between pH Levels and Target ANC Levels 

ANC Levels Related to Effects on Aquatic Ecosystems  

Consideration of Episodic Acidity

Consideration of Ecosystem Response Time 

Prior Examples of Target ANC Levels 

Consideration of Public Welfare Benefits

Summary of Alternative Levels

Combined Alternative Levels and Forms

Characterization of Uncertainties

Overview of Uncertainty

Uncertainties Associated with Data Gaps

Uncertainties in Modeled Processes

CASAC Advice

Administrator’s Proposed Conclusions

Field Pilot Program and Ambient Monitoring 

Field Pilot Program

Objectives

Overview of Field Pilot Program

Complementary Measurements

Complementary Research Efforts

Implementation Challenges

Final Monitoring Plan Development and Stakeholder Participation

Evaluation of Monitoring Methods

Potential FRMs for SO2 and p-SO4

Potential FRM for NOy	

Statutory and Executive Order Reviews

Executive Order 12866:  Regulatory Planning and Review

Paperwork Reduction Act

Regulatory Flexibility Act

Unfunded Mandates Reform Act

Executive Order 13132:  Federalism

Executive Order 13175:  Consultation and Coordination with Indian Tribal
Governments

Executive Order 13045:  Protection of Children From Environmental Health
and Safety Risks

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

National Technology Transfer and Advancement Act

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.”  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.”

	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 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 reviewed 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 Air Quality Criteria
for Oxides of Nitrogen was then finalized (US EPA, 1995a).  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.  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 (75 FR 6474), the secondary NAAQS for NO2 remains
0.053 ppm (100 micrograms per cubic meter [μg/m3] of air), annual
arithmetic average, calculated as the arithmetic mean of the 1-hour NO2
concentrations.

2.	NAAQS for Oxides of Sulfur

	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
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 Air Quality Criteria 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.  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.  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.  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 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, § 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.”


	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, EPA announced receipt of these
requests and asked for comment on the issues raised in them.  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 (65 FR 48699).

  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 Integrated
Science Assessment for Oxides of Nitrogen and Oxides of Sulfur-
Ecological Criteria (henceforth the "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.  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 a Risk and
Exposure Assessment for Review of the Secondary National Ambient Air
Quality Standards for Oxides of Nitrogen and Oxides of Sulfur
(henceforth 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 and public 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 Policy Assessment
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 Policy
Assessment (Russell and Samet, 2010a).   A second draft Policy
Assessment was released to CASAC and the public in September 2010 and
reviewed by CASAC on October 6-7, 2010.  CASAC provided advice and
recommendations to the Agency regarding the second draft Policy
Assessment in a December 9, 2010 letter (Russell and Samet 2010b). 
CASAC and public comments on the second draft Policy Assessment were
considered by EPA staff in developing a final Policy Assessment (US EPA,
2011).  CASAC requested an additional meeting to provide additional
advice to the Administrator based on the final Policy Assessment on
February 15–16, 2011.  EPA released on January 14, 2011, the final
Policy Assessment prior to final document production, to provide
sufficient time for CASAC review of the document in advance of this
meeting. A final Policy Assessment, 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
to the Administrator with regard to recommendations and revisions to the
secondary NAAQS for oxides of nitrogen and oxides of sulfur. 

alleging that EPA had failed to complete the current review within the
period provided by statute.  The schedule for completion of this review
is governed by a consent decree resolving that lawsuit and the
subsequent extension granted to the Agency by the plaintiffs 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.  

	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 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 N 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 Policy Assessment
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) that is based on the ecological criteria
described in the ISA and the results of the assessments in the REA, and
is 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.  Also consistent with the CAA, this
policy objective necessarily 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 Policy
Assessment (Russell and Samet, 2010a, 2010b), and as discussed in detail
in the Policy Assessment, 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, exposure, and ecological risks, presented more fully in the REA
and in the Policy Assessment, 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 Policy Assessment 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.   

. 

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 of the
ecological indicator for a targeted effect.

Drawing from the framework developed in the REA, the framework we used
to structure an ecologically meaningful secondary standard in the Policy
Assessment and to further develop the indicator, form, level, and
averaging time of such as 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.

Figure II-1. Simplified conceptual design of the form of an aquatic
acidification standard for oxides of nitrogen and sulfur.

The following discussion relies heavily on Chapters 2 and 3 of the
Policy Assessment.  The Policy Assessment 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 Policy Assessment 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 Policy Assessment 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 related to 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 both 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 the degree to which whether they are inadequate to protect
against such effects is presented.

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.  

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
sulfur dioxide (SO2), nitrogen dioxide (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).

Nature of ecosystem responses to gas-phase nitrogen and sulfur

	The 2008 ISA found that gas phase N and S 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. 

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 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.  

 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).

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
(SO42-) 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 waterbodies (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 NHx.

	When sulfur or nitrogen leaches from soils to surface waters in the
form of SO42- 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 calcium (Ca2+),
magnesium (Mg2+), sodium (Na+), or potassium (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
SO42- or NO3- leaching can deplete the available base cation pool in
soil. As the base cations are removed, continued deposition and leaching
of SO42- and/or NO3- (with H+ and Al3+) 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 N and S 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 was 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 United States. 
Acidification affects both terrestrial and freshwater aquatic
ecosystems.  

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.

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).  ANC is a key indicator
of acidification with relevance to both terrestrial and aquatic
ecosystems. 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), Ca2+, and aluminum. 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.  

. When ANC concentrations are <50 μeq/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 in 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).

, 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. 

	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 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/length3 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 µeq/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.  

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 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%, 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 Ca2+ 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
Ca2+ 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 S and N 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.

 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).

Magnitude of acidification-related ecosystem responses

	Terrestrial and aquatic ecosystems differ in their response to
acidifying 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.

Aquatic acidification

kes and streams (i.e., ANC less than about 50 μeq/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 μeq/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).

 	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 SO42- and NO3- concentrations and ANC
levels were analyzed to affirm the understanding that reductions in
deposition could influence the risk of acidification. 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)



Acute Concern	<0 μeq/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 μeq/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 μeq/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 μeq/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 μeq/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.



	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. MAGIC modeling predicts surface water concentrations of NO3-
and SO42- are10- 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 μeq/L (± 4.5 μeq/L). 55% of all monitored streams in the
Shenandoah Case Study Area have a current risk of Elevated, Severe, or
Acute.  Of the 55%, 18% 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 still have Elevated to Acute
problems with acidity in 2050. In fact, from 2006 to 2050, the
percentage of streams with Acute Concern is predicted to increase by 5%,
while the percentage of streams in Moderate Concern decreases by 5%.

sociated with decreasing stream ANC.  On average, the fish species
richness is lower by one fish species for every 21 μeq/L decrease in
ANC in Shenandoah National Park streams (US EPA, 2008, section 3.2.3.4).

 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 Ca2+, Mg2+ and potassium), as well as soil 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).	

 (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 United States 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%. Sugar
maple growth can be decreased by 20% 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 United States
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% to 75% of
all sugar maple plots across 24 states--that is, it exceeded the highest
(least stringent) of the three critical loads in 3% of plots, and the
lowest (most stringent) in 75% of plots. For red spruce, total nitrogen
and sulfur deposition levels exceeded three selected critical loads in
3% to 36% of all red spruce plots across eight states (US EPA, 2009,
section 4.4).  

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):

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.

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.

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.

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 REMAP surveys, provide considerable data on surface
water trends. 

.

ii.	Terrestrial acidification 

(US EPA, 2008, section 7.2.1 and Figure 7.2-1).  Sugar maple and red
spruce were the focus of the REA since they are demonstrated to be
negatively affected by soil available Ca2+ 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).

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 N greatly exceed N deposition.  Both oxides of
nitrogen and reduced forms of nitrogen (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.  

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.  

Aquatic ecosystems

− and dissolved inorganic nitrogen (DIN) concentration in surface
waters as well as Chl a:total P ratio. Elevated surface water NO3−
concentrations occur in both the eastern and western U.S. Studies report
a significant correlation between N deposition and lake biogeochemistry
by identifying a correlation between wet deposition and [DIN] and Chl a:
Total P. Recent evidence provides examples of lakes and streams that are
limited by N and show signs of eutrophication in response to N addition.

	The evidence is sufficient to infer a causal relationship between N
deposition and the alteration of species richness, species composition
and biodiversity in freshwater aquatic ecosystems (US EPA, 2008, section
3.3.5.3). Increased N 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 Nr deposition and the biogeochemical cycling of N and carbon (C)
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% (US EPA, 2009, section 5.1.1).

	The evidence is sufficient to infer a causal relationship between N
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 N contribute
to increased phytoplankton and algal productivity, leading to
eutrophication. Shifts in community composition, reduced hypolimnetic
DO, decreases in biodiversity, and mortality of submerged aquatic
vegetation are associated with increased N deposition in estuarine
systems. 

Terrestrial Ecosystems

	The evidence is sufficient to infer a causal relationship between N
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 indictors of N enrichment.  Several
N-addition studies indicate that NO3- leaching is induced by chronic
additions of N. Studies identified in the ISA found that surface water
NO3- concentrations exceeded 1 µeq/L in watersheds receiving about 9 to
13 kg N/ha/yr of atmospheric N deposition.  N 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 N mineralization, nitrification,
denitrification, foliar N concentration, and soil water NO3 − and NH4+
concentrations. Note that N saturation (N 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 N and C cycling, the
effects of N on C budgets (quantified input and output of C 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
N 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 N deposition on C budgets of non-forest
ecosystems.

	The evidence is sufficient to infer a causal relationship between N
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 N
deposition and effects of N deposition are not observed in all habitats.
 The most sensitive terrestrial taxa to N 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 N 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 N deposition. 

Ecosystem sensitivity to nutrient enrichment

	The numerous ecosystem types that occur across the U.S. have a broad
range of sensitivity to N deposition (US EPA, 2008, Table 4-4). 
Increased deposition to N-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 N, may result in an imbalance in
ecological stoichiometry, with effects on ecosystem processes, structure
and function. In general, N 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 N deposition levels as low as 3 kg N/ha/yr in
sensitive terrestrial ecosystems to N deposition. In aquatic ecosystems,
N 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 N loading
and N-sensitivity.  N-sensitivity is predominately driven by the degree
to which growth is limited by nitrogen availability. Grasslands in the
western United States are typically N-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 N
fertilization increased aboveground growth in all non-forest ecosystems
except for deserts. In other words, almost all terrestrial ecosystems
are N-limited and will be altered by the addition of anthropogenic
nitrogen. Likewise, a freshwater lake or stream must be N-limited to be
sensitive to N-mediated eutrophication. There are many examples of fresh
waters that are N-limited or N 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 N-limitation occurred as
frequently as P-limitation in freshwater ecosystems.   Additional
factors that govern the sensitivity of ecosystems to nutrient enrichment
from N deposition include rates and form of N deposition, elevation,
climate, species composition, plant growth rate, length of growing
season, and soil N 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 N deposition compared to acidification.

	Because the productivity of estuarine and near shore marine ecosystems
is generally limited by the availability of N, they are susceptible to
the eutrophication effect of N 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.  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). 

the somewhat arbitrary discreteness of the EI scale can mask the
benefits of decreases in nitrogen between categories.

	In general, estuaries tend to be N-limited, and many currently receive
high levels of N input from human activities to cause eutrophication. As
reported in the ISA (US EPA, 2008, section 3.2.2.2), atmospheric N loads
to estuaries in the U.S. are estimated to range from 2-8% for Guadalupe
Bay, TX on the lowest end to as high as 72% for St Catherines-Sapelo
estuary, GA. The Chesapeake Bay is an example of a large, well-studied
and severely eutrophic estuary that is calculated to receive as much as
30% of its total N load from the atmosphere.

Magnitude of ecosystem responses

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 N deposition.  As noted previously, studies
found that N-limitation occurs as frequently as P-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 N deposition show increased
inorganic N concentration and productivity to be correlated with
atmospheric N deposition. The results are unequivocal evidence of N
limitation in lakes with low ambient inputs of N, and increased N
concentrations in lakes receiving N solely from atmospheric N
deposition. It has been suggested that most lakes in the northern
hemisphere may have originally been N-limited, and that atmospheric N
deposition has changed the balance of N and P in lakes.

	Available data suggest that the increases in total N 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 N 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 N 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
N deposition was estimated by to range from 0.4 to 0.7 kg N/ha/yr.

−, 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 N load
to estuaries contributes to the total load but does not necessarily
drive the effects since other combined sources of N often greatly exceed
N deposition.  In estuaries, N-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 United States.

	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. 

.  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.

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 N deposition. As
previously stated, most terrestrial ecosystems are N-limited, therefore
they are sensitive to perturbation caused by N additions (US EPA, 2008,
section 4.3.1). Effects are most likely to occur where areas of
relatively high atmospheric N deposition intersect with N-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 United States are being altered by elevated levels of N inputs,
including inputs from atmospheric deposition.

	In the eastern U.S., the degree of N 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 stages of saturation
across several regions in the eastern U.S. Of the 85 northeastern
watersheds examined 60% were in Stage 1 or Stage 2 of N 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 Policy
Assessment.  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 NADP monitoring data.  This
map, reproduced in Figure 3-5 of the Policy Assessment, depicts the
sites where empirical effects of terrestrial nutrient enrichment have
been observed and site proximity to elevated atmospheric N 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 CCS 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’s 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:

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

10 kg N/ha/yr – mycorrhizal community changes

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

3.1 kg N/ha/yr – shift from sensitive to tolerant lichen species

5.2 kg N/ha/yr – dominance of the tolerant lichen species

10.2 kg N/ha/yr – loss of sensitive lichen species

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% 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. 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. 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.

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. 

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).

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. 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
methylmercury can be produced in the absence of sulfate. Once
methylmercury is present, other variables influence how much accumulates
in fish, but elevated mercury levels in fish can only occur where
substantial amounts of methylmercury are present. Current evidence
indicates that in watersheds where mercury is present, increased oxides
of sulfur deposition very likely results in additional production of
methylmercury which leads to greater accumulation of MeHg concentrations
in fish. With respect to sulfur deposition and mercury methylation, the
final ISA determined: The 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 methylmercury (MeHg) requires
the presence of SO42- and mercury, and where mercury is present,
increased availability of SO42- 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 methylmercury requires
the presence of sulfate and mercury, but the amount of methylmercury
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. 

	While the relationship between sulfur and methylmercury production was
concluded to be causal in the ISA, the REA concluded that there was
insufficient evidence to quantify the relationship between sulfur and
methylmercury.  Therefore only a qualitative assessment was included in
Chapter 6 of the REA. The Policy Assessment 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 United States.  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
ecosystem responses adverse effects associated with adverse effects
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
maintain the functioning of 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 as well as evaluating
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 National Ambient Air Quality Standards (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
methylmercury production; nitrous oxide (N2O) effects on climate;
nitrogen effects on primary productivity and biogenic greenhouse gas
fluxes; and phytotoxic effects on plants. 

	Because the targeted ecosystem effects outlined above are not evenly
distributed across the United States, 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, WI) 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 Coastal Sage Scrub
Communities in southern California and the Mixed Confer Forest
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.	

Air quality analyses

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 National Atmospheric Deposition Program
(NADP) National Trends Network and dry deposition predictions from the
2002 Community Multi-Scale Air Quality (CMAQ) model simulation.  Some
key conclusions from this analysis are:

Total reactive nitrogen deposition and sulfur deposition are much
greater in the East compared to most areas of the West. 

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.

Reduced nitrogen deposition exceeds oxidized nitrogen deposition in the
vicinity of local sources of NH3.

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.

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 Mixed Conifer Forest Case Study
Area, in which sulfur deposition is greatest in winter.

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 SO42-and NO3-concentrations and acid neutralizing capacity
(ANC) levels and to affirm the understanding that reductions in
deposition could influence the risk of acidification.  Monitoring data
from the EPA-administered Temporally Integrated Monitoring of Ecosystems
(TIME)/Long-Term Monitoring (LTM) programs and the Environmental
Monitoring and Assessment Program (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. 

valents per liter (μeq/L) (± 15.7 μeq/L); 78 % of all monitored lakes
in the Adirondack Case Study Area have a current risk of Elevated,
Severe, or Acute. Of the 78%, 31% experience episodic acidification, and
18% are chronically acidic today.

Based on the steady-state critical load model for the year 2002, 18%,
28%, 44%, and 58% of 169 modeled lakes received combined total sulfur
and nitrogen deposition that exceeded their critical loads, with
critical corresponding to ANC limits of 0, 20, 50, and 100 μeq/L
respectively.

 concentrations in surface waters. ANC levels increased from about 50
μeq/L in the early 1990s to >75 μeq/L until 2002, when ANC levels
declined back to 1991–1992 levels. Current concentrations are still
above pre-acidification (1860) conditions. MAGIC modeling predicts
surface water concentrations of NO3 and SO42- are 10- and 32-fold higher
today, respectively. The estimated average ANC for 60 streams in the
Shenandoah Case Study Area is 57.9 μeq/L (± 4.5 μeq/L). 55% of all
monitored streams in the Shenandoah Case Study Area have a current risk
of Elevated, Severe, or Acute. Of the 55%, 18% experience episodic
acidification, and 18% are chronically acidic today.

 ANC limits of 0, 20, 50, and 100 μeq/L respectively.

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 United States by using the base cation to aluminum (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%. Sugar
maple growth can be reduced by 20% at a Bc/Al ratio of 0.6.  Key
findings of the case study are summarized below.

eq/ha/yr for the Bc/Al ratios of 0.6, 1.2, and 10.0 (increasing levels
of tree protection).

.

.

Deposition-related aquatic nutrient enrichment

A summary of findings follows:

2002 CMAQ/NADP results showed that an estimated 40,770,000 kg of total
nitrogen was deposited in the Potomac River watershed. SPARROW modeling
predicted that 7,380,000 kg N/yr of the deposited nitrogen reached the
estuary (20% 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). 

 that a decrease of at least 78% in the 2002 total nitrogen atmospheric
deposition load to the watershed would be required.

2002 CMAQ/NADP results showed that an estimated 18,340,000 kg of total
nitrogen was deposited in the Neuse River watershed. SPARROW modeling
predicted that 1,150,000 kg N/yr of the deposited nitrogen reached the
estuary (26% of the total load to the estuary). The overall ASSETS EI
for the Neuse River/Neuse River Estuary was Bad.

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.

. A waterbody’s response to nutrient loading depends on the magnitude
(e.g., agricultural sources have a high influence in the Neuse), spatial
distribution, and other characteristics of the sources within the
watershed.

Deposition-related terrestrial nutrient enrichment

	California Coastal Sage Scrub (CSS) and Mixed Conifer Forest (MCF)
communities were the focus of the Terrestrial Nutrient Enrichment Case
Studies of the REA. 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’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:

 Although CSS decline has been observed in the absence of fire, the
contributions of deposition and fire to the CSS decline require further
research. 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.

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 United States
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
particulate matter (PM) NAAQS review currently underway.  A discussion
of the causal relationship between sulfur deposition (as sulfate, SO42-)
and increased mercury methylation in wetlands and aquatic environments
is also included in the REA. On this subject the REA concludes that
decreases in sulfate deposition will likely result in decreases in
methyl mercury concentration; however, spatial and biogeochemical
variations nationally hinder establishing large scale dose-response
relationships.  

The heterogeneity of ecosystems across the United States, however,
introduces variations into dose-response relationships.

	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, that further assessment was not warranted at this
time. 

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 United States has not been quantified.

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 Clean Air Act, 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 related to impacts on the public from
alterations in structure and function of ecosystems is an appropriate
consideration for this review.  

	Based on information provided in the Policy Assessment, the following
section discusses how ecological effects from deposition of oxides of
nitrogen and sulfur relate to adversity to public welfare.  In the
Policy Assessment, 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, including potential value of the services.  The quantified
indicators of value discussed below relate to ecosystem services
generally that are impacted by deposition related acidification, not to
the specific effects of this action. 

Ecosystem Services   

	The Policy Assessment 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 Policy Assessment 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. 

	As 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 Policy Assessment 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
given ecological indicator (e.g., for aquatic acidification, to changes
in acid neutralizing capacity [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).  	 

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. 

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 United States 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 Policy Assessment, 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% 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 United States 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 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 Adiriondack case study to all 3,000
Adirondack lakes and by applying the willingness to pay 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 µeq/L.  The REA
estimated 44% of the Adirondack lakes currently fall below an ANC of 50
µeq/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
µeq/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 overall impact
on these services is likely to be substantial.

Terrestrial Acidification

	Chapter 4.4.3 and 4.4.4 of the Policy Assessment review several
economic studies of areas sensitive to terrestrial acidification. 
Forests in the northeastern United States 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 Policy Assessment, 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 United States.   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.

Nutrient Enrichment

	Chapter 4.4.5 and 4.4.6 of the Policy Assessment 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% 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 Policy Assessment,
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 acid 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 Policy Assessment reviewed studies related to park use in
California.  Data from California State Parks indicate that in 2002,
68.7% of adult residents participated in trail hiking for an average of
24.1 days per year. The analyses in the Policy Assessment 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 acid deposition, or of any
specific reductions in deposition, relative to the effects of many other
factors that may affect them.

	The Policy Assessment also identified fire regulation as a service that
could be affected by nutrient over-enrichment of the Coastal Sage Scrub
and Mixed Conifer Forest 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.

Summary

	Adversity to public welfare can be understood not only by looking at
how deposition of oxides of nitrogen and sulfur affect the ecological
functions of an ecosystem (see II.A.), but also and then by
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 acid 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.  Important biodiversity is 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 and nitrogen and sulfur standard 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 Policy Assessment, 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 theis general
conclusion that aquatic and terrestrial ecosystems can be degrade by
deposition of oxides of nitrogen and sulfur, it is then necessary to
first evaluate the appropriateness 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.  

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, e.g. 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.  

Appropriateness and Adequacy of the Current Standards for
Deposition-related Effects

	This section addresses two concepts necessary to evaluate the current
standards on 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
relates evaluates the current standards and in terms of adequacy of
protection.  

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 N and S 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 N and S deposition.  	

 issues arise that call into question the ecological relevance of the
structure of the current secondary standards for oxides of nitrogen and
sulfur.  

 standards do not utilize appropriate atmospheric indicators.  NO2 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.  

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 N
and S from both atmospheric deposition of oxides of nitrogen and sulfur
as well as the inputs from other sources of N and S 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.  Although the standards do not
directly address reduced forms of nitrogen in the atmosphere, e.g. they
do not require specific levels of reduced nitrogen, iIt 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 N
and S.

Based on the discussioned summarized above, the Policy Assessment
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.

Adequacy of Protection

 In addition, these levels based on conclusions in the REA 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 Policy Assessment 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 indicates that atmospheric N deposition is the main source of
new anthropogenic N to most headwater streams, high elevation lakes, and
low-order streams. Atmospheric N deposition contributes to the total N
load in terrestrial, wetland, freshwater, and estuarine ecosystems that
receive N 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 S
deposition in the US.  For the period 2004–2006, mean S 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 S
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. S 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 SO42−.  		

 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 are as most
relevant to evaluating the adequacy of the existing standards in
protecting public welfare from adverse ecological effects.

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 IIB 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 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. 

 This information indicates that almost half of the 44 lakes in the
Adirondacks case study area are at an elevated concern levels, 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 μeq/L relative
to an ANC level at 100 μeq/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 μeq/L and 13 percent of lakes exceeded the critical load for
an ANC of 20 μeq/L.

As with the Adirondacks area, this information suggests that significant
numbers of sensitive streams in the Shenandoah area are at risk of
adverse impacts on fish populations under recent conditions. 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 Policy Assessment all conclude that the current
standards are not adequate to protect against the adverse impacts of
aquatic acidification on sensitive ecosystems.  In the ISA it is noted
that significant portions of the U.S. are acid sensitive, and that
current deposition levels exceed those that would allow 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). 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 IIB).  Based on these
considerations, the Policy Assessment 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.  

Terrestrial acidification

	Based on the terrestrial acidification case studies (Kane Experimental
Forest in Pennsylvania and Hubbard Brook Experimental Forest (case study
locations described in section IIB) on sugar maple and red spruce
habitat, the REA concludes that there is significant risk to sensitive
terrestrial ecosystems from acidification at atmospheric concentrations
of NOx and SOX at or below the current standards.    The ecological
indicator selected for terrestrial acidification is the base cation to
aluminum ratio (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.)
and many 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).  

.  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.

	

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 United States
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 NOx 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 N-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 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 the12 km, 2002 CMAQ/NADP data, with
the exception of the easternmost Sierra Nevadas. 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. 

Aquatic nutrient enrichment	

.  In addition, this type of indicator does not reflect the impact of
nitrogen deposition in conjunction with other sources of nitrogen.  

	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 N nutrient
enrichment effects in estuarine aquatic ecosystems.

Other effects

neurotoxic contaminant.  The production of meaningful amounts of
methylmercury (MeHg) requires the presence of SO42- and mercury, and
where mercury is present, increased availability of SO42- 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 methylmercury requires the presence of sulfate and
mercury, but the amount of methylmercury 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: The 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 Policy Assessment 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 Policy Assessment
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 Policy Assessment 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.

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 Policy Assessment 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.”

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 IIA. The
Administrator also recognizes the need for conclusions both 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 Policy Assessment, 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, annual average concentration,
measured in the ambient air as NO2, and the current secondary standard
for oxides of sulfur at 0.5 ppm, 3-hour average concentration, measured
in the ambient air as SO2. 

	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 Policy
Assessment 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 Policy Assessment, 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 to 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 Policy Assessment 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
Policy Assessment would employ (1) 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 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, termed an aquatic acidification
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 Policy
Assessment 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 Policy Assessment also
considers how measurable or quantifiable an indicator is to enable its
use as an effective indicator of relevant ambient air concentrations.

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 Policy Assessment 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

 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 Policy Assessment is
based on monitoring or modeling of these species.  The Policy Assessment
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.

2.	Oxides of Nitrogen

	As discussed in the Policy Assessment (US EPA, 2011,section 7.1.2),
total reactive oxidized nitrogen, NOy, as defined in chapter 2 of the
Policy Assessment, 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.  NOy is
an aggregate measure of s 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.  NOy
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 Policy Assessment, 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 Policy Assessment.  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 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 Policy Assessment 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 Policy
Assessment.  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 Policy Assessment 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 Policy Assessment 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 Policy Assessment, 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 nitrogen dioxide and,
consequently, for a similar ambient air concentration, HNO3 contributes
more deposition of acidifying nitrogen relative to nitrogen dioxide.  In
transitioning from source-oriented urban locations to rural
environments, the ratio of the concentrations of HNO3 and PAN to NO2
increases.

 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).

.

B.	Form

	Based on the evidence of the aquatic acidification effects caused by
the deposition of NOy and SOx, the Policy Assessment (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 Policy
Assessment 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 this 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 Policy Assessment for
selecting acid neutralizing capacity (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:

ANC is directly associated with the causative agents, pH and dissolved
aluminum, both through empirical evidence and mechanistic relationships;

Empirical evidence shows very clear and strong relationships between
adverse effects and ANC;

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

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 Policy Assessment, atmospheric
deposition of NOy and SOx causes aquatic acidification through the input
of strong acid anions (e.g., NO3- and SO42-) 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 acid neutralizing capacity (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 (SO42-, NO3-), cations (Ca2+, Mg+2, K+),
or sums of cations or anions.

	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 to 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.   ANC is preferred
over ALK as the body of scientific evidence has focused on ANC and
effects relationships.   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 corollary
coverse of ANC from the perspective of 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, aluminum, 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 Policy Assessment 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.

	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.   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.   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 Policy Assessment
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 Policy Assessment (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 Policy Assessment.  In large measure, this
dependence between pH 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
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 Policy Assessment 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.  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 other 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 Policy Assessment 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 µeq/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 µeq/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 Policy Assessment 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 Policy
Assessment notes that in its review of the first draft Policy
Assessment, 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 Policy Assessment, 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 Model of Acidification of Groundwater in
Catchment (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 Policy Assessment.

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 (US 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 ueq/L in 1990 to 67 ueq/L in 2006 (US 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 (US EPA, 2011, section 2
and US 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 Policy Assessment, 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 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 Policy Assessment 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 Policy Assessment 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 Policy Assessment 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 Policy Assessment
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.  It is expressed as:

CLANClim(N + S) = ([BC]0* - [ANClim])Q + Neco					(III-1)

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 Policy Assessment.  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 Policy
Assessment.  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.  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. 

	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 Policy Assessment 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., m3 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 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.   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 Policy Assessment
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 Policy Assessment, 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:

CLANClim(N + S) = ([BC]0* - [ANClim])Q + Neco = [NOy]TNOy + [SOx]TSOx +
NHx	(III-2)

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, oxidize 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 (US 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 Policy Assessment 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 Policy Assessment 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 Policy Assessment. 

	As discussed in Appendix F, the interannual variability, as well as the
sensitivity to emission changes of roughly 50%, results in changes of
transference ratios of approximately 5 - 10%.  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 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 Policy Assessment concludes
that the transference ratios are relatively stable and provide a sound
metric for linking deposition and concentration in the form of the
standard.

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 Policy Assessment (US EPA,
2011,section 2) considers CMAQ and other models, such as CAMx and the
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 for purposes of developing
a new secondary standard for consideration.  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, is supported by a central infrastructure of EPA
scientists, with considerable interfacing with the scientific research
communities in academia and industry,  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
1980’s of specific improvements in deposition processes, which are
described in Appendix F of the Policy Assessment.

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 Policy Assessment derives the following expression of these
linkages, which separates reduced forms of nitrogen, NHx, from oxidized
forms:

 - TNOy [NOy]/Qr - TSOx[SOx]/Qr 		(III-3)

 

Based on equation III- 3, the Policy Assessment defines an aquatic
acidification index (AAI) that is more simply stated using terms that
highlight the ambient air indicators:

AAI  =  F1 – F2 – F3[NOy] – F4[SOx]						(III- 4)

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 ;

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]}, being greater than Neco.  In
rare cases where Neco is greater than {NHx + F3[NOy]}, F2, are 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 Policy Assessment concludes that equation III- 4 (US 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 association
with 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 Policy Assessment, 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 Policy Assessment 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 Policy Assessment (US 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 United States.  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 Policy Assessment
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 Policy Assessment 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 Policy
Assessment concludes that Omernik’s ecoregion classification 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 Policy Assessment concludes that ecoregion level III
(Figure III-1) resolution, with 84 defined ecoregions in the contiguous
U.S., 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 Policy Assessment concludes that the most detailed
level of resolution (level IV) is not appropriate given the limited data
availability to address nearly 1000 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.

	Appendix C of the Policy Assessment includes a description of each
level III ecoregion.  The Policy Assessment notes that the use of
ecoregions is an appropriate spatial aggregation scheme for this
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 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.

Figure III-1.  Omernik Ecoregion III areas ( HYPERLINK
"http://www.epa.gov/wed/pages/ecoregions"
http://www.epa.gov/wed/pages/ecoregions ).

a.	Ecoregion Sensitivity

	The Policy Assessment used Omernik’s original alkalinity data (US
EPA, 2011,as 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.   	

 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,  HYPERLINK
"http://www.epa.gov/mrlc/nlcd-2006.html"
http://www.epa.gov/mrlc/nlcd-2006.html ), acid sensitive ecoregions
would meet both of the following land use data criteria: percent of
developed and agricultural area less than 20% combined with forested
area greater than 50%.  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.  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.

Figure III-2.  Acid-sensitive ecoregions identified in grey fill (22 out
of 84 ecoregions).

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 Policy Assessment 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 Policy
Assessment 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 of calculated critical loads was created for the water
bodies in each ecoregion for which sufficient water quality and
hydrology data are available.  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.

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

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. 
Also 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 Policy Assessment 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 related deposition, while potentially affording
little or no public welfare benefit from attempting to improve a
naturally acidic system.

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 Policy
Assessment 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 Policy Assessment, 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 Policy Assessment 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 Policy
Assessment notes that modeled NHx deposition exhibits greater spatial
variability than the other modeled terms in factors F3 and F4. 
Recognizing this greater variability, the Policy Assessment 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 Policy Assessment 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 Policy Assessment 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 Policy Assessment 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 Policy Assessment
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 Policy Assessment 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 Policy Assessment 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 these area 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 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 Policy Assessment 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 Policy
Assessment 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 Policy Assessment 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 Policy
Assessment 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 acid neutralizing capacity (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 Policy Assessment developed such a form, termed an aquatic
acidification index (AAI), 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
Policy Assessment 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 Policy Assessment 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 Policy 	Assessment 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 Policy Assessment 	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 rations 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 Policy Assessment 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 Policy Assessment.     

C.	Averaging Time

	As discussed in section 7.3 of the Policy Assessment, 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 Policy
Assessment 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 Policy Assessment 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 Policy Assessment 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 Policy Assessment 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
Policy Assessment 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 Policy Assessment notes that in its
comments on the second draft Policy Assessment, CASAC  agreed that a 3
to 5 year averaging time was appropriate to consider (Russell and Samet,
2010b).

D.	Level

	As discussed above, the Policy Assessment 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.  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 Policy Assessment 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.   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.

	The Policy Assessment 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 Policy Assessment
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 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
Policy Assessment, 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. 


y causal indicator of effects related to aquatic acidification, this
suggests that ANC values below approximately -50 μeq/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 μeq/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 Policy
Assessment was on target ANC values in the range of -50 to 100 μeq/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. 

ly -50 to 100 μeq/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 μeq/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
μeq/L), severe concern (0-20 μeq/L), elevated concern (20-50 μeq/L),
moderate concern (50-100 μeq/L), and low concern (>100 μeq/L).  This
categorization is supported by a large body of research completed
throughout the eastern U.S. (Sullivan et al., 2006).

Policy Assessment has focused on target ANC levels no lower than 0
μeq/L.

	As discussed in the Policy Assessment, biota generally are not harmed
when ANC values are >100 μeq/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 μeq/L.  This suggests that at ANC levels
greater than 100 μeq/L, little risk from acidification exists in many
aquatic ecosystems.  At ANC levels below 100 μeq/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 μeq/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 μeq/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 Policy Assessment concludes that the above considerations
support a focus on target ANC levels up to a level greater than 50
μeq/L but below 100 μeq/L, such as up to a level of 75 μeq/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 Policy Assessment, CASAC expressed the following views
about the range of biological responses that corresponds to this range
of ANC levels (i.e., 0-100 μeq/L):

ects at ANC levels below 20 μeq/L, and reasonable confidence that there
are adverse effects below 50 μeq/L.  Levels of 50 μeq/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 μeq/L. (Russell and Samet, 2010b)

 

	The Policy Assessment 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 μeq/L.

3.	Consideration of Episodic Acidity  

	As discussed in the Policy Assessment, across the broad range of ANC
values from 0 to 100 μeq/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
Policy Assessment.  Accordingly, greater protection would be achieved by
target chronic ANC values set high enough to avoid pH depression to
levels associated with elevated risk.   

	The specific relationship between ANC and the probability of reaching
pH levels of elevated risk varies by water body and fish species.  ANC
levels below 20 μeq/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 μeq/L.  Overall, there is little
uncertainty that significant effects on aquatic biota are occurring at
ANC levels below 20 μeq/L.  

	It is clear that at ANC levels approaching 0 μeq/L (Table II-1), there
is significant impairment of sensitive aquatic ecosystems with almost
complete loss of fish species.  Avoiding ANC levels approaching 0 μeq/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 Policy Assessment
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 μeq/L.  Staddard et al. (2003) noted on average a 30 μeq/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 μeq/L.  The above considerations do not provide support
for a target chronic ANC level as low as 0 μeq/L for a standard that
would protect against significant harm to aquatic ecosystems, including
harm from episodic acidification.  The Policy Assessment concludes that
these considerations also support a lower end of the range for
consideration no lower than 20 μeq/L.

	The CASAC agreed with this conclusion in its comments on the second
draft Policy Assessment (Russell and Samet, 2010b).  CASAC noted that
“there are clear and marked biological effects at ANC values near 0
μeq/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:

] approximately 50 μeq/L.  Thus, based on these studies, a long term
ANC target level of 75 μeq/L would generally guard against effects from
episodic acidification down to a level of about 25 μeq/L. (Russell and
Samet, 2010b)

4.	Consideration of Ecosystem Response Time  

	The Policy Assessment 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 (US 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
μeq/L.  Therefore, for these water bodies, reductions in acidifying
deposition are not likely to achieve an ANC of 50 μeq/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.

el of the ecological indicator by a given time.  For example, to achieve
an ANC level of 20 μeq/L by 2030, it might be necessary to specify a
higher target ANC level of, for example, 50 μeq/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
μeq/L as a long-term equilibrium level.  In this example, the target
ANC of 50 μeq/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
Policy Assessment 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 Policy Assessment 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 Policy Assessment 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 Policy
Assessment, critical loads were established to protect 95 percent of
surface waters in Europe from an ANC less than 20 µeq/L based on
protection of brown trout.  Individual countries have set alternative
ANC targets; for example, Norway targets an ANC of 30 µeq/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
µeq/L or possibly higher.  New Hampshire and Vermont have set ANC
targets of 60 µeq/L and 50 µeq/L, respectively.  Tennessee has
established site-specific target ANC values based on assessments of
natural acidity, with a default value of 50 µeq/L when specific data
are not available.

hat target ANC values between 20 and 60 μeq/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)).

While the text above lists a number of welfare effects, the NAAQS is
aimed at protection from adverse effects to public welfare. 
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 Policy Assessment.  

, the study implies benefits to the New York population roughly on the
order of $600 million per year (in constant 2007$).  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 Policy Assessment 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 μeq/L may be adverse to public
welfare.

7.	Summary of Alternative Levels

	Based on all the above considerations, the Policy Assessment concludes
that consideration should be given to a range of standard levels from 20
to 75 μeq/L.  The available evidence indicates that target ANC levels
below 20 μeq/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 μeq/L in some sensitive ecosystems, the degree and
nature of those effects are less significant than at levels below 20
μeq/L.  Levels at and above 50 μeq/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 μeq/L, as effects are
generally appreciably less sensitive to changes in ANC at such higher
levels.

	In reaching this conclusion in the Policy Assessment, 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 Policy Assessment also note 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 Policy Assessment 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 Policy Assessment 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 µeq/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 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.

 In all cases, these relatively non-acid sensitive ecoregions were
estimated to meet all of the alternative standards considered in this
assessment.

	As described above, the AAI values presented in Table 7-1a-d of the
Policy Assessment 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 (US 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 µeq/L, 90th percentile) to 4 for
the least protective alternative standard (20 µeq/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 Policy Assessment 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 (US 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 Policy Assessment 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 Policy Assessment 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 Policy Assessment 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 Policy Assessment 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% and 42%, respectively,
relative to 2005 base year emissions.  The Policy Assessment 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 µeq/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 (US
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 and 48% 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 anthropogenic
contributions of mineral acids is likely responsible for observed low
ANC values in those regions.

	The Policy Assessment notes the Sierra Nevada region as an interesting
case study, as it has some of the lowest critical load values nationally
(US 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 Policy
Assessment.  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 Policy Assessment 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 Policy
Assessment, 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 Policy Assessment concludes that the behavior
of the AAI calculations is both reasonable and explainable, which the
Policy Assessment 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 Policy Assessment (US 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 Policy Assessment (US 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
(US 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 anioni and cation species as defined in equation 2.11 in
the Policy Assessment), 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 reduced nitrogen (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
not point sources), there have been more frequent deposition flux
studies, relative to other nitrogen species, that enable both 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 Policy Assessment 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 deposition.  Also, from an aquatic acidification perspective, NHx is
not as important a driver relative to 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 Policy Assessment
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 Policy Assessment 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 National Parks Service (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 National Atmospheric Deposition Program’s
(NADP) passive ammonia network (AMoN).  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 s
representativeness of modeled NHx deposition within and across
ecoregions.

	With regard to water quality data, the Policy Assessment 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 Policy Assessment) 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 Policy Assessment, 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 Integrated Science Assessment 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
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 Policy Assessment 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-b ased standard, an improved
understanding 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 Policy Assessment 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 Policy Assessment 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 Integrated Science Assessment,
the Risk and Exposure Assessment, and the Policy Assessment.  CASAC has
endorsed EPA’s interpretation of the science embodied in the
Integrated Science Assessment and the assessment approaches and
conclusions incorporated in the Risk and Exposure Assessment.

	Most recently, CASAC has considered the information in the final Policy
Assessment 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
Policy Assessment 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)

	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 additional,
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 total reactive oxidized nitrogen (NOy) and sulfur oxides
(SOx) as atmospheric  indicators of oxidized nitrogen (N) and sulfur (S)
atmospheric concentrations is well justified.  The use in the Aquatic
Acidification Index (AAI) of NOy and SOx as atmospheric indicators of N
and S concentrations is useful and corresponds with other efforts by
EPA. As we have stated previously, CASAC also agrees that acid
neutralizing capacity (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 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:

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
μeq/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)

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
(μeq/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.  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 Policy Assessment,
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 Policy
Assessment, 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

As discussed above, the Administrator notes that the Integrated Science
Assessment 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 Integrated Science
Assessment 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 Policy Assessment, and
supported by CASAC, that there is a strong scientific basis for
development of a standard with the general structure presented in the
Policy Assessment.

	The Administrator also recognizes that the conceptual framework for an
ecologically relevant, multi-pollutant standard, which was initially
explored in the Risk and Exposure Assessment and further developed in
the Policy Assessment, builds on the information in the Integrated
Science Assessment.  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.

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
determine an appropriate form an level for such a standard, within the
ranges of levels and forms identified in the Policy Assessment, and the
representativeness of F factors in the AAI equation described above and
in the Policy Assessment.  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
Policy Assessment, 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 the CAA.

	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
Policy Assessment.  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 tp 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.  The issue here is
identifying a specific form and level for the standard that would
provide an appropriate level of public welfare policy judgment to be
made protection, consistent with the requirements of the CAA, in light
of the specific uncertainties and limitations that are discussed in
detail throughout this preamble.  As noted above, it 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 form and level for the 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 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 the
any particular intended degree of protection of public welfare that the
Administrator determined satisfied the requirements of the CAA.  	While
acknowledging that CASAC supported moving forward to establish the
standard developed in the Policy Assessment, the Administrator also
observes that CASAC supported conducting further field studies that
would better inform the continued development or modification of such as
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 Policy Assessment.  This will provide a better basis for the
Administrator to identify a form and level of the standard that is
requisite to protect the public welfare consistent with the CAA,
specifically with respect to the acidifying effects of deposition of
oxides of nitrogen and sulfur.  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.

add secondary standards of the ecologically relevant form to address
acid depositions, 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 necessary 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 proposed Transport 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 oxides of
nitrogen (NOy) and oxides of sulfur (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
Policy Assessment.  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.  EPA notes
that the monitoring program described here is intended to supplement
existing monitoring programs and will require additional resources.

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 five-year period in
a sample of 3 to 5 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 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, and to support the
identification of a specific form and level for such a standard that
would be requisite to protect public welfare consistent with the
requirements 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 oxides of nitrogen and sulfur,
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 oxides of nitrogen; and

(7)  support related longer-term research efforts, including
enhancements to and evaluation of modeled dry deposition algorithms.

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. 
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 five-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 verify
and enhance the performance of the established methods, data retrieval
and reporting procedures.  This would in turn inform the future
selection of specific parameters for, and application of the
measurements into the AAI equation.

Figure IV-1.  Location of CASTNET sites in relation to acid sensitive
ecoregions. 

   

Over the course of this five-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 an 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 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 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, 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: by evaluating total NOy instrument behavior, and air quality
model evaluation.  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 five-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 suggests 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.  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 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?

(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 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 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 3-5
ecoregions and the instrumentation to be deployed.  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 ORD
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 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 oxides of nitrogen (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 are suitable approaches 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 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  HYPERLINK
"http://www.epa.gov/castnet" 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® filter
that collects p-SO42- and p-NO3, the second stage contains a nylon
filter that collects SO2 (as SO42-) and HNO3, and the third stage
contains two cellulose fiber filters impregnated with potassium
carbonate (K2CO3) that collect any remaining SO2 (as SO42-).  The
sampler collects one-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% – 10% (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 National
Park Service (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 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 µg/m3, 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 µg/m3 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 further evaluate 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 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® 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) × 1.03 – 0.42; r2
= 0.9933) and night time ([NOy](ppb) = [NOyi](ppb) × 1.01 – 0.18; r2
= 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.  NOy instruments have been
routinely operated in networks such as 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.

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.” because it was deemed
to “raise novel legal or policy issues.”  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 Unfundated 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 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.  However, aAs 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.

	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

 action is not subject to EO 13045 because it is not an economically
significant rule as defined in EO 12866.  addresses the effects on
public welfare of oxides of nitrogen and sulfur and does not establish
an environmental standard intended to mitigate health or safety risks.

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 No. 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.

  EPA is not aware of any voluntary consensus standards that are
relevant to the provisions of this proposed rule.  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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Authority (NYSERDA).

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 Ambient
Air Quality Standards for Nitrogen Dioxide and Sulfur Dioxide. U.S.
Environmental Protection Agency, Research Triangle Park, NC,
EPA-452/R-08-006.

US EPA 2008. Integrated Science Assessment (ISA) for Oxides of Nitrogen
and Sulfur Ecological Criteria (Final Report). U.S. Environmental
Protection Agency, Washington, D.C., EPA/600/R-08/082F, 2008.

US EPA 2009. Risk and Exposure Assessment for Review of the 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-09-008a

US EPA, 2010a. CASTNET Quality Assurance Project Plan, Revision 7.0,
October 2010,  HYPERLINK "http://java.epa.gov/castnet/"
http://java.epa.gov/castnet/ 

US EPA, 2010b.  CASTNET Annual Reports, 2004 - 2009,  HYPERLINK
"http://java.epa.gov/castnet/" http://java.epa.gov/castnet/ 

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, D.C.,
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.

Wolff, G. T. 1993. CASAC closure letter for the 1993 Criteria Document
for Oxides of Nitrogen addressed to U.S. EPA Administrator Carol M.
Browner dated September 30, 1993.

Wolff, G. T. 1995. CASAC closure letter for the 1995 OAQPS Staff Paper
addressed to U.S. EPA Administrator Carol M. Browner dated August 22,
1995.

List of Subjects in 40 CFR Part 50

	Environmental protection, Air pollution control, Carbon monoxide, Lead,
Nitrogen dioxide, Ozone, Particulater 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:

§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 amended by deleting paragraphs (f) and (g) and
revising paragraphs (a) through (e) to read as follows: 

§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:

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 §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) The terms used in this appendix are defined as follows: 

*     *     *     *     * 

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
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) 

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 Sulfur as measured
by Sulfur Dioxide (“SO2 NAAQS”) specified in § 50.17 and §50.5
(b), respectively, are met at an ambient air quality monitoring site. 
Sulfur Dioxide (SO2) is measured in the ambient air by a Federal
reference method (FRM) based on appendix A-1 or A-2 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 SO2 concentrations and the levels of
the SO2 NAAQS are specified in the following sections.

*     *     *     *     *

(c) The terms used in this appendix are defined as follows: 

*     *     *     *     *

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 value for the primary and
secondary 1-hour NAAQS is the 3-year average of annual 99th percentile
daily maximum 1-hour values for a monitoring site (referred to as the
“1-hour primary standard design value”).

*     *     *     *     *

2. Requirements for Data Used for Comparisons with the SO2 NAAQS and
Data

Reporting Considerations.

*     *     *     *     *

(b) Data from two or more monitors from the same year at the same site
reported to EPA under distinct Pollutant Occurrence Codes shall not be
combined in an attempt to meet data completeness requirements.  The
Administrator will combine annual 99th percentile daily maximum
concentration values from different monitors in different years,
selected as described here, for the purpose of developing a valid 1-hour
primary or secondary standard design value.  If more than one of the
monitors meets the completeness requirement for all four quarters of a
year, the steps specified in section 5(a) of this appendix shall be
applied to the data from the monitor with the highest average of the
four quarterly completeness values to derive a valid annual 99th
percentile daily maximum concentration.  If no monitor is complete for
all four quarters in a year, the steps specified in section 3(c) and
5(a) of this appendix shall be applied to the data from the monitor with
the highest average of the four quarterly completeness values in an
attempt to derive a valid annual 99th percentile daily maximum
concentration.  This paragraph does not prohibit a monitoring agency
from making a local designation of one physical monitor as the primary
monitor for a Pollutant Occurrence Code and substituting the 1-hour data
from a second physical monitor whenever a valid concentration value is
not obtained from the primary monitor; if a monitoring agency
substitutes data in this manner, each substituted value must be
accompanied by an AQS qualifier code indicating that substitution with a
value from a second physical monitor has taken place.

*     *     *     *     *

3. Comparisons with the 1-hour Primary and Secondary SO2 NAAQS.

 (a) The 1-hour primary or secondary SO2 NAAQS is met at an ambient air
quality monitoring site when the valid 1-hour primary or secondary
standard design value is less than or equal to 75 parts per billion
(ppb).

(b) An SO2 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(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.

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 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.

 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).

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

 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.

 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).

 As discussed in chapter 2 of the Policy Assessment, SO2 and particulate
SO4 are routinely measured in ambient air monitoring networks, although
only 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.

 The Policy Assessment 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.

  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.

 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.

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

 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.

 The Policy Assessment judged the data to be sufficient for this purpose
if data are available from more than 10 water bodies in an ecoregion.

 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.

 Tables 7-1a-d and 7-2 in the Policy Assessment 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.

Interagency Working Comments on Draft Language under EO12866 and 13563
Interagency Review.  Subject to Further Policy Review. 

  PAGE   \* MERGEFORMAT  1 

Clarify that you are also keeping the current secondary standards.

I routinely suggest editing this out of boilerplate for water rules.  I
think it sounds unprofessional to say this to the public.  OW has not
had any problem accommodating this suggestion.

There seem to be two closed quotes for a single open quote.  Is this two
quotes or one?  Please fix.

Please clarify what original complaint was amended.

Are these defined terms of art or are you just using these distinctions
casually?  If the former, might be helpful to explain the technical
distinction among the three levels of confidence causality in a
footnote.

Not sure if there is a typo here or if phototoxic and phytotoxic are two
different terms that are being used deliberately.

In t he previous sentence, you say that fish fitness begins to decline
at 100.  Here you say that fitness of sensitive species begins to
decline between 100 and 50.  Not sure if this is redundant or what
distinction you are trying to make.

Would be helpful to add a footnote explaining how ANC is measured
(conceptually, not physically) and what it means to have an ANC < 0.

Wouldn’t the ecological effect result from the interaction of ANC and
load?  That is, if you had low ANC but also very low loadings, you would
not necessarily see the effects cited here, or do I misunderstand? 
Would help to clarify this someplace.

This notation may be confusing to some readers.  Suggest you use
eq/ha/yr consistently throughout (as you do elsewhere).

Not clear what this means.  What is “pre-acidification?”  What
“difference” are you referring to here?

Not sure what this means.

I thought the exact quantification of this function was one of the areas
in which there was substantial uncertainty, which is part of your basis
for deferring setting of the ecologically relevant standard.  Please
clarify.

I don’t understand what you are saying here.  Do you mean that
empirically N and S are linked so that neither can be zero.  Or do you
mean that the model constrains both, or the combination of the two, to
be non-zero.  Please clarify.

Not clear from this sentence what was compared to what to get this 20%
decrease.

Unless you have a basis for believing that controls on atmospheric
deposition would be less costly per unit of nitrogen than controls on
other sources (no such evidence is presented here), suggest deleting
this speculative statement.  Does not fit with the generally rigorous
scientific tone of this discussion.

Not sure what you are saying here.  Do you mean that these 65% of
estuaries had the greatest loads relative to other estuaries, relative
to other types of water bodies, or are you saying something about
relative contribution from different source types?

What are non-ambient loadings?

This partial statistic is not very informative.  Would be useful to know
what happens to the entire range of concern (presumably elevated to
acute) collectively, as well as how this overall change breaks down
between the three categories elevated, severe, and acute.

Doesn’t the lowest critical load go with the highest level of
protection?  If so, I would say, “…ranged from 6,008 to 107
eq/ha/yr…” If I understand how this works, a low critical load gives
you a high Bc/Al level.

See prev comment.

This is confusing.  I thought critical load was tied to a specific
target ANC level.  If that is now how you defined critical load here,
would be helpful to clarify.

This is a pretty vague statement. Do you mean greater in some areas and
close in others.  Also I assume that this depends on the critical load
chosen.  Since the loads studies vary by 1.5 orders of magnitude, I
assume the relationships to actual deposition will vary a lot depending
on which load you are using.  It would be helpful to make this statement
more specific.

Do you have any reason to think that deposition levels will rise? 
Don’t you have regs recently enacted or in the pipeline that are
likely to significantly reduce NOx/Sox emissions.  Suggest sticking to
the model results and not speculating about what might happen in the
future.

Aren’t these both estuaries where N deposition is known to contribute
significantly to total loadings?  Suggest deleting this speculative
sentence.

Not clear what this means.  Do you mean a slim chance that it would be
enough, or a slim chance that this much would be needed.  Also, if this
comes from a quantified estimate, suggest replacing “slim” with a
more precise statement (preferably quantitative).

I note that the Neuse has a higher share of N coming from deposition
than the Potomac, but that eliminating 78% of deposition could improve
the Potomac from Bad to Poor, while decreasing deposition by any amount
apparently could not produce this result in the Neuse.  It would be
helpful to provide additional explanation for this apparent anomaly.

How do you know this.  This statement appears to be undercut by the next
two sentences.  Please clarify.

It is confusing to have a range of inputs associated with a single point
estimate for emissions changes.  Please clarify what exactly was modeled
and what it showed.

Suggest breaking number one into two as noted below and saying five
issues.  Bullet one here addresses two apparently unrelated issues,
averaging time and appropriate indicators.

Suggest beginning bullet 2 here.  From here to the end of the para
treats a separate issue.

This issue is not unique to NOx/SOx.  Suggest requesting comment on
potential implementation challenges associated with a standard that
varies by region, since a single emissions source may impact multiple
regions.  Modeling to figure out necessary reductions will be
complicated and could become intractable.

Are these levels lower than in the past?  If so, this pattern could also
suggest that there are long lag times in recovery.

This info is distracting and not needed here, also these statistic are
about absolute levels of deposition, not proportions as mentioned
earlier in the sentence.  Suggest deleting.

This insert is suggested so as not to imply that the earlier standards
were not important.

Is this equivalent to saying that 44 percent had an ANC below 50 and 28%
had an ANC below 20?  If so, wouldn’t this be a simpler way to say it,
and also more consistent with preceding and following sentences?  If
not, is the distinction between being below a specified ANC level, and
exceeding the critical load for an ANC level significant?  May be
helpful to clarify.

These two sentences seem inconsistent, unless exceeding the critical
load for an ANC of 50 is not equivalent to having an ANC below 50 (see
previous comment).  Please clarify.

Para above suggests that 0.6 is the relevant ratio for sugar maple.

Again, para above identifies a critical load of 1.2 for red spruce.

Not much context for this statement.  Critical loads based on what ANC
level.  Is 250 a lot?  What would the impact on ANC be of an exceedence
of this magnitude?

This does not seem like a very strong basis for concluding that the
current standards are inadequate, but I’m not sure if this is your
purpose here or not.  Again, do you have any reason to believe that N
and S emission levels will rise in the future.  I would have thought the
opposite.

N/ha/yr

This only shows that these ecosystems are receiving N above the level
necessary for healthy plant growth.  How does it follow that deposition
above this level is adverse?  If N is not limiting above this point,
wouldn’t addtl N deposits have no effect, either positive or negative?
 Please clarify.

Are such change significant.  If so, why does the index only have six
levels?

Statement as worded seems strong for the evidence provided.  How about
suggested edits?

My toxicologist suggests that it is more scientifically neutral to
simply say “neurotoxic.”  How neurotoxic in any given circumstance
is dose-dependent.  Also, she notes that mercury vapor is much more
neurtoxic.

Is this preamble the first place where you have formally articulated
this conclusion:?  Is this a policy or a scientific conclusion?

Why only oxides of N?  Why is NOy separated out from other sources of N
that can also lead to acid deposition?  Wouldn’t a standard that
reflected all such sources of N make more sense?

Why is this not included in the list in the previous sentence?

Do you mean “despite”?

Given that the different components of NOy have different deposition
velocities, what single velocity would you use to represent NOy globally
in calculating the specific standard?

Is this a word? Neither I nor spell check recognize it.  What does it
mean?

Why is there no T coefficient for this term?

I spent a long time puzzling over this equation.  I think this sentence
would help explain what you are doing here.

These seem like very stringent criteria for identifying acid sensitive
regions, given that the table II-1 identifies any ANC level above 100 as
being of “low concern.”  (200 is not even identified as a threshold
of concern in the table.) You are saying that if even 1% of water bodies
exceed this low concern threshold, the region is “acid sensitive.” 
What is the basis for these cut points?  They do not seem to fit well
with the info in Table II-1. 

Would be helpful to clarify that when you say 70th to 90th percentile,
you are looking at the CLs arranged from highest to lowest.  If they are
arranged in the more intuitive ascending order, this would actually 10th
to 30th percentile, or do I misunderstand what you are doing?

Lowest = most difficult to achieve, no?

I don’t understand the logic here.  If you set at median, you protect
only 50% of water bodies.  If you set at 90th percentile, you protect 90
percent of water bodies.  What does it matter what share of water bodies
in a region are “acid sensitive.”  This is a continuum, not a bright
line.  In fact, by using a lower percentile in non-acid sensitive
regions, you are actually protecting a far lower share of any acid
sensitive water bodes that happen to be in that region.  Consider the
following thought experiment.  If 50% of the water bodies in a region
are acid sensitive, these will presumably correspond to the upper half
of the distribution, and protecting at the 90th percentile level will
protect 80% of these acid sensitive water bodies (plus all of the
non-acid sensitive ones).  On the other hand, if only 20% of the water
bodies are acid sensitive (a less acid sensitive region) then protecting
at the 90th percentile will protect only half of this smaller number of
acid sensitive water bodies.  This problem is further exacerbated if you
instead select the median.  In this case you would not be protecting any
acid sensitive water bodies in this region.

If you use the same percentile in both cases, you will still
(appropriately) get a more stringent standard in acid sensitive regions
(bc 90th percentile in acid sensitive region is more protective than
90th percentile in non-acid sensitive regions) but you will be
protecting the same share of water bodies in both cases. To me this
makes more sense.

This is only true if there are not enough unimpaired lakes to satisfy
current fishing demand.  If there are, adding more will have little
economic benefit (though still significant ecological benefits).  The
information presented here provides no basis for determining if this is
true or not.

Not clear what you are saying here.  Are you saying that a modeled
increment of improvement from 20 to 50 has $600 million in benefits. 
This statement is hard to evaluate if the starting point for all the
lakes is not 20 (presumably there is a wide variation in baseline ANCs).
 Alternately, you could mean that cleaning up all the lakes to somewhere
in the range of 20 to 50 provides these benefits, but then I don’t see
how you get a point estimate for benefits.  Wouldn’t cleaning up to 50
give much higher benefits than cleaning up to 20?  Please explain.

It would be helpful to include in this preamble a table showing the
calculated AAI values for all eco-regions, including both acid sensitive
and non-acid sensitive.

What averaging time are you suggesting using in the standard.  I thought
it was annual, or is it longer?

On p 10 you say “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.” I suggest adapting and
emphasizing this language in this section because it is an important
basis for deciding not to set an ecologically relevant standard now. 
Although you are confident in the basic form, you do not have enough
info to set a standard that is “neither more nor less stringent than
necessary.”  It is not good enough under the statute to simply err in
the direction of making the standard overly protective, as some may urge
you to do.  It is exactly this need to “get it right” that prevents
you from setting a standard now.  

I think it is important to emphasize that this is more than just a
discretionary policy choice, which some may argue you have no authority
to make, having established that the current standard is insufficient to
protect public welfare.  As I understand it, the reason you are not
setting a standard at this time is that even though you have enough info
to determine the current standard is inadequate, you do not have enough
information to establish one that better complies with the new science. 
You are confident in the structure of this better standard, but do not
have the data to establish its specific form and level.  Thus, this is
not just a policy choice.  Rather, it is not feasible at this time to
set a standard that meets the statutory objective of being requisite to
protect public welfare without being either more or less stringent than
necessary.

I would be happy to discuss this further by phone on Monday.

Don’t think you want to say the current secondary standards are
inappropriate.  They are appropriate to address direct effects of
gaseous NOx/SOx.

How do you know this?  Please explain the basis for this statement.

We do not agree with this interpretation, which is not in the EO.  In
this case, it is clear that the rule is not subject to the EO because it
is not an economically significant rule, one prong of the two-pronged
applicability test in the EO itself.

I don’t believe this is accurate.  How about the proposed alternative.

