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 review of the oxides of nitrogen
secondary standards in 1987.  In November 1991 EPA released an updated
draft air quality criteria document (AQCD) for CASAC and public review
and comment (56 FR 59285), which provided a comprehensive assessment of
the available scientific and technical information on health and welfare
effects associated with NO2 and other oxides of nitrogen.  The CASAC
reviewed the draft document at a meeting held on July 1, 1993 and
concluded in a closure letter to the Administrator that the document
“provides a scientifically balanced and defensible summary of current
knowledge of the effects of this pollutant and provides an adequate
basis for EPA to make a decision as to the appropriate NAAQS for NO2”
(Wolff, 1993).  The 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. 

	 On February 3, 2006, the Center for Biological Diversity and four
other plaintiffs filed an amended complaint 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. 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 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

	Freshwater ecosystem surveys and monitoring in the eastern United
States have been conducted by many programs since the mid-1980s,
including EPA’s Environmental Monitoring and Assessment Program
(EMAP), National Surface Water Survey (NSWS), Temporally Integrated
Monitoring of Ecosystems (TIME), and Long-term Monitoring (LTM)
programs. Based on analyses of surface water data from these programs,
New England, the Adirondack Mountains, the Appalachian Mountains
(northern Appalachian Plateau and Ridge/Blue Ridge region), and the
Upper Midwest contain the most sensitive lakes and streams (i.e., ANC
less than about 50 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.

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 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).

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.

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?

