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6560-50-P

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.

SUMMARY: This proposed rule is being issued as required by a consent decree governing the 

schedule for completion of this review of the air quality criteria and the secondary national 

ambient air quality standards (NAAQS) for oxides of nitrogen and oxides of sulfur. Based on its 

review, EPA proposes to retain the current nitrogen dioxide (NO2) and sulfur dioxide (SO2) 

secondary standards to provide requisite protection for the direct effects on vegetation resulting 

from exposure to gaseous oxides of nitrogen and sulfur in the ambient air. Additionally, with 

regard to protection from the deposition of oxides of nitrogen and sulfur to sensitive aquatic and 

terrestrial ecosystems, including acidification and nutrient enrichment effects, EPA is proposing 

to add secondary standards identical to the NO2 and SO2 primary 1-hour standards and not set a 

new multi-pollutant secondary standard in this review. The proposed 1-hour secondary NO2

standard would be set at a level of 100 ppb and the proposed 1-hour secondary SO2 standard 

would be set at 75 ppb. In addition, EPA has decided to undertake a field pilot program to gather 

and analyze additional relevant data so as to enhance the Agency’s understanding of the degree 

of protectiveness that a new multi-pollutant approach, defined in terms of an aquatic acidification 
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index (AAI), would afford and to support development of an appropriate monitoring network for 

such a standard. The EPA solicits comment on the framework of such a standard and on the 

design of the field pilot program. The EPA will sign a notice of final rulemaking for this review 

no later than March 20, 2012.

DATES: Written comments on this proposed rule must be received by [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:

 www.regulations.gov: Follow the on-line instructions for submitting comments.

 Email: 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.
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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 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 www.regulations.gov or 

email. The 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 

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 

http://www.epa.gov/epahome/dockets.htm. 

Docket: All documents in the docket are listed in the 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 www.regulations.gov or in hard copy at the Air and Radiation 
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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?

1. Submitting CBI. Do not submit this information to EPA through 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.

2. Tips for Preparing Your Comments. When submitting comments, remember to:
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 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.

 Explain your views as clearly as possible.

 Make sure to submit your comments by the comment period deadline identified.

Availability of Related Information

A number of documents relevant to this rulemaking are available on EPA web sites. The 

Integrated Science Assessment for Oxides of Nitrogen and Sulfur - Ecological Criteria: Final 

Report (ISA) is available on EPAs National Center for Environmental Assessment web site. To 

obtain this document, go to http://www.epa.gov/ncea, and click on Air Quality then click on 

Oxides of Nitrogen and Sulfur. The Policy Assessment (PA), Risk and Exposure Assessment

(REA), and other related technical documents are available on EPA’s Office of Air Quality

Planning and Standards (OAQPS) Technology Transfer Network (TTN) web site. The PA is 

available at http://www.epa.gov/ttn/naaqs/standards/no2so2sec/cr_pa.html, and the exposure and 
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risk assessments and other related technical documents are available at 

http://www.epa.gov/ttn/naaqs/standards/no2so2sec/cr_rea.html. These and other related 

documents are also available for inspection and copying in the EPA docket identified above.

Table of Contents

The following topics are discussed in this preamble:

I. Background

A. Legislative Requirements

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

1. NAAQS for Oxides of Nitrogen

2. NAAQS for Oxides of Sulfur

C. History of Related Assessments and Agency Actions

D. History of the Current Review

E. Scope of the Current Review

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

A. Ecological Effects

1. Effects Associated with Gas-Phase Oxides of Nitrogen and Sulfur

a. Nature of ecosystem responses to gas-phase nitrogen and sulfur

b. Magnitude of ecosystem response to gas-phase nitrogen and sulfur

2. Acidification Effects Associated with Deposition of Oxides of Nitrogen 

and Sulfur

a. Nature of Acidification –related Ecosystem Responses

i. Aquatic ecosystems

ii. Terrestrial ecosystems

iii. Ecosystem sensitivity

b. Magnitude of Acidification-related Ecosystem Responses

i. Aquatic acidification

ii. Terrestrial acidification

c. Key Uncertainties Associated with Acidification

i. Aquatic acidification

ii. Terrestrial acidification

3. Nutrient Enrichment Effects Associated with Deposition of Oxides of 

Nitrogen 

a. Nature of Nutrient Enrichment-related Ecosystem Responses

i. Aquatic ecosystems

ii. Terrestrial ecosystems

iii. Ecosystem sensitivity to nutrient enrichment

b. Magnitude of Nutrient Enrichment-related Ecosystem Responses

i. Aquatic ecosystems

ii. Terrestrial ecosystems

c. Key Uncertainties Associated with Nutrient Enrichment

i. Aquatic ecosystems
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ii. Terrestrial ecosystems

4. Other Ecological Effects

B. Risk and Exposure Assessment

1. Overview of Risk and Exposure Assessment

2. Key Findings

a. Air quality analyses

b. Deposition-related aquatic acidification

c. Deposition-related terrestrial acidification

d. Deposition-related aquatic nutrient enrichment

e. Deposition-related terrestrial nutrient enrichment 

f. Additional effects

3. Conclusions on Effects

C. Adversity of Effects to Public Welfare

1. Ecosystem Services

2. Effects on Ecosystem Services

a. Aquatic Acidification

b. Terrestrial Acidification

c. Nutrient Enrichment

3. Summary

D. Adequacy of the Current Standards

1. Adequacy of the Current Standards for Direct Effects

2. Appropriateness and Adequacy of the Current Standards for Depositionrelated Effects

a. Appropriateness

b. Adequacy of Protection

i. Aquatic Acidification

ii. Terrestrial acidification

iii. Terrestrial nutrient enrichment

iv. Aquatic nutrient enrichment

v. Other effects

3. CASAC Views

4. Administrator’s Proposed Conclusions Concerning Adequacy of Current 

Standard

III. Rationale for Proposed Decision on Alternative Multi-pollutant Approach to Secondary

Standards for Aquatic Acidification

A. Ambient Air Indicators

1. Oxides of Sulfur

2. Oxides of Nitrogen

B. Form

1. Ecological Indicator

2. Linking ANC to Deposition

3. Linking Deposition to Ambient Air Indicators

4. Aquatic Acidification Index

5. Spatial Aggregation

a. Ecoregion Sensitivity

b. Representative Ecoregion-specific Factors
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i. Factor F1

a) Acid-sensitive Ecoregions

b) Non-acid sensitive Ecoregions

ii. Factor F2

iii. Factors F3 and F4

c. Factors in Data-limited Ecoregions

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

6. Summary of the AAI Form

C. Averaging Time

D. Level

1. Association Between pH Levels and Target ANC Levels 

2. ANC Levels Related to Effects on Aquatic Ecosystems 

3. Consideration of Episodic Acidity

4. Consideration of Ecosystem Response Time 

5. Prior Examples of Target ANC Levels 

6. Consideration of Public Welfare Benefits

7. Summary of Alternative Levels

E. Combined Alternative Levels and Forms

F. Characterization of Uncertainties

1. Overview of Uncertainty

2. Uncertainties Associated with Data Gaps

3. Uncertainties in Modeled Processes

4. Applying Knowledge of Uncertainties

G. CASAC Advice

H. Administrator’s Proposed Conclusions

IV. Field Pilot Program and Ambient Monitoring 

A. Field Pilot Program

1. Objectives

2. Overview of Field Pilot Program

3. Complementary Measurements

4. Complementary Areas of Research Implementation Challenges

5. Final Monitoring Plan Development and Stakeholder Participation

B. Evaluation of Monitoring Methods

1. Potential FRMs for SO2 and p-SO4

2. Potential FRM for NOy

V. Statutory and Executive Order Reviews

A. Executive Order 12866: Regulatory Planning and Review

B. Paperwork Reduction Act

C. Regulatory Flexibility Act

D. Unfunded Mandates Reform Act

E. Executive Order 13132: Federalism

F. Executive Order 13175: Consultation and Coordination with Indian Tribal 

Governments

G. Executive Order 13045: Protection of Children From Environmental Health and 

Safety Risks
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H. Executive Order 13211: Actions that Significantly Affect Energy Supply, 

Distribution, or Use

I. National Technology Transfer and Advancement Act

J. Executive Order 12898: Federal Actions to Address Environmental Justice in 

Minority Populations and Low-Income Populations

References

I. Background

A. Legislative Requirements

Two sections of the Clean Air Act (CAA) govern the establishment and revision of the 

NAAQS. Section 108 (42 U.S.C. section 7408) directs the Administrator to identify and list 

certain air pollutants and then to issue air quality criteria for those pollutants. The Administrator 

is to list those air pollutants that in her “judgment, cause or contribute to air pollution which may

reasonably be anticipated to endanger public health or welfare;” “the presence of which in the 

ambient air results from numerous or diverse mobile or stationary sources;” and “for which . . . 

[the Administrator] plans to issue air quality criteria...” Air quality criteria are intended to 

“accurately reflect the latest scientific knowledge useful in indicating the kind and extent of all 

identifiable effects on public health or welfare which may be expected from the presence of [a] 

pollutant in the ambient air . . .” 42 U.S.C. § 7408(b). Section 109 (42 U.S.C. 7409) directs the 

Administrator to propose and promulgate “primary” and “secondary” NAAQS for pollutants for 

which air quality criteria are issued. Section 109(b)(1) defines a primary standard as one “the 

attainment and maintenance of which in the judgment of the Administrator, based on such 

criteria and allowing an adequate margin of safety, are requisite to protect the public health.”1 A 

secondary standard, as defined in section 109(b)(2), must “specify a level of air quality the 

attainment and maintenance of which, in the judgment of the Administrator, based on such 

 

1 The legislative history of section 109 indicates that a primary standard is to be set at “the maximum permissible 

ambient air level . . . which will protect the health of any [sensitive] group of the population,” and that for this 

purpose “reference should be made to a representative sample of persons comprising the sensitive group rather than 

to a single person in such a group” S. Rep. No. 91-1196, 91st Cong., 2d Sess. 10 (1970).
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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
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1. NAAQS for Oxides of Nitrogen

After reviewing the relevant science on the public health and welfare effects associated 

with oxides of nitrogen, EPA promulgated identical primary and secondary NAAQS for NO2 in 

April 1971. These standards were set at a level of 0.053 parts per million (ppm) as an annual 

average (36 FR 8186). In 1982, EPA published Air Quality Criteria Document for Oxides of 

Nitrogen (US EPA, 1982), which updated the scientific criteria upon which the initial standards 

were based. In February 1984 EPA proposed to retain these standards (49 FR 6866). After 

taking into account public comments, EPA published the final decision to retain these standards 

in June 1985 (50 FR 25532).

The EPA began the most recent previous review of the oxides of nitrogen secondary

standards in 1987. In November 1991, EPA released an updated draft air quality criteria 

document (AQCD) for CASAC and public review and comment (56 FR 59285), which provided 

a comprehensive assessment of the available scientific and technical information on health and 

welfare effects associated with NO2 and other oxides of nitrogen. The CASAC reviewed the 

draft document at a meeting held on July 1, 1993 and concluded in a closure letter to the 

Administrator that the document “provides a scientifically balanced and defensible summary of 

current knowledge of the effects of this pollutant and provides an adequate basis for EPA to 

make a decision as to the appropriate NAAQS for NO2” (Wolff, 1993). The AQCD for Oxides 

of Nitrogen was then finalized (US EPA, 1995a). The EPA’s OAQPS also prepared a Staff 

Paper that summarized and integrated the key studies and scientific evidence contained in the 

revised AQCD for oxides of nitrogen and identified the critical elements to be considered in the 

review of the NO2 NAAQS. The CASAC reviewed two drafts of the Staff Paper and concluded 
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in a closure letter to the Administrator that the document provided a “scientifically adequate 

basis for regulatory decisions on nitrogen dioxide” (Wolff, 1995).

In October 1995, the Administrator announced her proposed decision not to revise either 

the primary or secondary NAAQS for NO2 (60 FR 52874; October 11, 1995). A year later, the 

Administrator made a final determination not to revise the NAAQS for NO2 after careful 

evaluation of the comments received on the proposal (61 FR 52852; October 8, 1996). While the 

primary NO2 standard was revised in January 2010 by supplementing the existing annual 

standard with the establishment of a new 1-hour standard, set at a level of 100 ppb (75 FR 6474), 

the secondary NAAQS for NO2 remains 0.053 ppm (100 micrograms per cubic meter [μg/m3] of 

air), annual arithmetic average, calculated as the arithmetic mean of the 1-hour NO2 

concentrations.

2. The NAAQS for Oxides of Sulfur

The EPA promulgated primary and secondary NAAQS for SO2 in April 1971 (36 FR 

8186). The secondary standards included a standard set at 0.02 ppm, annual arithmetic mean, 

and a 3-hour average standard set at 0.5 ppm, not to be exceeded more than once per year. These 

secondary standards were established solely on the basis of evidence of adverse effects on 

vegetation. In 1973, revisions made to Chapter 5 (“Effects of Sulfur Oxide in the Atmosphere on 

Vegetation”) of the AQCD for Sulfur Oxides (US EPA, 1973) indicated that it could not properly

be concluded that the vegetation injury reported resulted from the average SO2 exposure over the 

growing season, rather than from short-term peak concentrations. Therefore, EPA proposed (38 

FR 11355) and then finalized (38 FR 25678) a revocation of the annual mean secondary

standard. At that time, EPA was aware that then-current concentrations of oxides of sulfur in the 

ambient air had other public welfare effects, including effects on materials, visibility, soils, and 
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water. However, the available data were considered insufficient to establish a quantitative 

relationship between specific ambient concentrations of oxides of sulfur and such public welfare 

effects (38 FR 25679).

In 1979, EPA announced that it was revising the AQCD for oxides of sulfur concurrently 

with that for particulate matter (PM) and would produce a combined PM and oxides of sulfur 

criteria document. Following its review of a draft revised criteria document in August 1980, 

CASAC concluded that acid deposition was a topic of extreme scientific complexity because of 

the difficulty in establishing firm quantitative relationships among (1) emissions of relevant 

pollutants (e.g., SO2 and oxides of nitrogen), (2) formation of acidic wet and dry deposition 

products, and (3) effects on terrestrial and aquatic ecosystems. The CASAC also noted that acid 

deposition involves, at a minimum, several different criteria pollutants: oxides of sulfur, oxides 

of nitrogen, and the fine particulate fraction of suspended particles. The CASAC felt that any

document on this subject should address both wet and dry deposition, since dry deposition was 

believed to account for a substantial portion of the total acid deposition problem.

For these reasons, CASAC recommended that a separate, comprehensive document on 

acid deposition be prepared prior to any consideration of using the NAAQS as a regulatory

mechanism for the control of acid deposition. The CASAC also suggested that a discussion of 

acid deposition be included in the AQCDs for oxides of nitrogen and PM and oxides of sulfur. 

Following CASAC closure on the AQCD for oxides of sulfur in December 1981, EPA’s OAQPS 

published a Staff Paper in November 1982, although the paper did not directly assess the issue of 

acid deposition. Instead, EPA subsequently prepared the following documents to address acid 

deposition: The Acidic Deposition Phenomenon and Its Effects: Critical Assessment Review 

Papers, Volumes I and II (US EPA, 1984a, b) and The Acidic Deposition Phenomenon and Its 
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Effects: Critical Assessment Document (US EPA, 1985) (53 FR 14935 -14936). These 

documents, though they were not considered criteria documents and did not undergo CASAC 

review, represented the most comprehensive summary of scientific information relevant to acid 

deposition completed by EPA at that point.

In April 1988 (53 FR 14926), EPA proposed not to revise the existing primary and 

secondary standards for SO2. This proposed decision with regard to the secondary SO2 NAAQS 

was due to the Administrator’s conclusions that: (1) based upon the then-current scientific 

understanding of the acid deposition problem, it would be premature and unwise to prescribe any

regulatory control program at that time; and (2) when the fundamental scientific uncertainties 

had been decreased through ongoing research efforts, EPA would draft and support an 

appropriate set of control measures. Although EPA revised the primary SO2 standard in June 

2010 by establishing a new 1-hour standard at a level of 75 ppb and revoking the existing 24-

hour and annual standards (75 FR 35520), no further decisions on the secondary SO2 standard 

have been published. 

C. History of Related Assessments and Agency Actions

In 1980, the Congress created the National Acid Precipitation Assessment Program 

(NAPAP) in response to growing concern about acidic deposition. The NAPAP was given a 

broad 10-year mandate to examine the causes and effects of acidic deposition and to explore 

alternative control options to alleviate acidic deposition and its effects. During the course of the 

program, the NAPAP issued a series of publicly available interim reports prior to the completion 

of a final report in 1990 (NAPAP, 1990).

In spite of the complexities and significant remaining uncertainties associated with the 

acid deposition problem, it soon became clear that a program to address acid deposition was 
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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 
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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.
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In a related matter, the Office of the Secretary of the U.S. Department of Interior (DOI) 

requested in 2000 that EPA initiate a rulemaking proceeding to enhance the air quality in 

national parks and wilderness areas in order to protect resources and values that are being 

adversely affected by air pollution. Included among the effects of concern identified in the 

request were the acidification of streams, surface waters, and/or soils; eutrophication of coastal 

waters; visibility impairment; and foliar injury from ozone.

In a Federal Register notice in 2001 (65 FR 48699), EPA announced receipt of these 

requests and asked for comment on the issues raised in them. The EPA stated that it would 

consider any relevant comments and information submitted, along with the information provided 

by the petitioners and DOI, before making any decision concerning a response to these requests 

for rulemaking.

The 2005 NAPAP report states that “... scientific studies indicate that the emission 

reductions achieved by Title IV are not sufficient to allow recovery of acid-sensitive ecosystems. 

Estimates from the literature of the scope of additional emission reductions that are necessary in 

order to protect acid-sensitive ecosystems range from approximately 40-80% beyond full 

implementation of Title IV....” The results of the modeling presented in this Report to Congress 

indicate that broader recovery is not predicted without additional emission reductions (NAPAP, 

2005).

Given the state of the science as described in the ISA, REA, and in other recent reports, 

such as the NAPAP reports noted above, EPA has decided, in the context of evaluating the 

adequacy of the current NO2 and SO2 secondary standards in this review, to revisit the question 

of the appropriateness of setting secondary NAAQS to address remaining known or anticipated 
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adverse public welfare effects resulting from the acidic and nutrient deposition of these criteria 

pollutants.

D. History of the Current Review

The EPA initiated this current review in December 2005 with a call for information (70 

FR 73236) for the development of a revised ISA. An Integrated Review Plan (IRP) was 

developed to provide the framework and schedule as well as the scope of the review and to 

identify policy-relevant questions to be addressed in the components of the review. The IRP was 

released in 2007 (US EPA, 2007) for CASAC and public review. The EPA held a workshop in 

July 2007 on the ISA to obtain broad input from the relevant scientific communities. This 

workshop helped to inform the preparation of the first draft ISA, which was released for CASAC 

and public review in December 2007; a CASAC meeting was held on April 2-3, 2008 to review 

the first draft ISA. A second draft ISA was released for CASAC and public review in August 

2008, and was discussed at a CASAC meeting held on October 1-2, 2008. The final ISA (US 

EPA, 2008) was released in December 2008. 

Based on the science presented in the ISA, EPA developed the REA to further assess the 

national impact of the effects documented in the ISA. The Draft Scope and Methods Plan for 

Risk/Exposure Assessment: Secondary NAAQS Review for Oxides of Nitrogen and Oxides of 

Sulfur outlining the scope and design of the future REA was prepared for CASAC consultation

and public review in March 2008. A first draft REA was presented to CASAC and the public for 

review in August 2008 and a second draft was presented for review in June 2009. The final REA 

(US EPA, 2009) was released in September 2009. A first draft PA was released in March 2010 

and reviewed by CASAC on April 1-2, 2010. In a June 22, 2010 letter to the Administrator, 

CASAC provided advice and recommendations to the Agency concerning the first draft PA
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(Russell and Samet, 2010a). A second draft PA was released to CASAC and the public in 

September 2010 and reviewed by CASAC on October 6-7, 2010. The CASAC provided advice 

and recommendations to the Agency regarding the second draft PA in a December 9, 2010 letter 

(Russell and Samet 2010b). The CASAC and public comments on the second draft PA were 

considered by EPA staff in developing a final PA (US EPA, 2011). CASAC requested an 

additional meeting to provide additional advice to the Administrator based on the final PA on 

February 15–16, 2011. On January 14, 2011, EPA released a version of the final PA prior to 

final document production, to provide sufficient time for CASAC review of the document in 

advance of this meeting. The final PA, incorporating final reference checks and document 

formatting, was released in February 2011.. In a May 17, 2011 letter (Russell and Samet, 2011a),

CASAC offered additional advice and recommendations to the Administrator with regard to the

review of the secondary NAAQS for oxides of nitrogen and oxides of sulfur. 

In 2005, the Center for Biological Diversity and four other plaintiffs filed a complaint 

alleging that EPA had failed to complete the current review within the period provided by 

statute.2 The schedule for completion of this review is governed by a consent decree resolving 

that lawsuit and the subsequent extension agreed to by the parties. The schedule presented in the 

original consent decree that governs this review, entered by the court on November 19, 2007, 

was revised on October 22, 2009 to allow for a 17-month extension of the schedule. The current 

decree provides that EPA sign for publication notices of proposed and final rulemaking 

concerning its review of the oxides of nitrogen and oxides of sulfur NAAQS no later than July

12, 2011 and March 20, 2012, respectively. 

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 

 

2 Center for Biological Diversity, et al. v. Johnson, No. 05–1814 (D.D.C.)
Page 20 of 238

of conclusions, findings, and determinations proposed by the Administrator are noted. While 

they identify the reasoning that supports this proposal, they are only proposals and are not 

intended to be final or conclusive in nature. The EPA invites general, specific, and/or technical 

comments on all issues involved with this proposal, including all such proposed judgments, 

conclusions, findings, and determinations.

E. Scope of the Current Review

In conducting this periodic review of the secondary NAAQS for oxides of nitrogen and 

oxides of sulfur, as discussed in the IRP and REA, EPA decided to assess the scientific 

information, associated risks, and standards relevant to protecting the public welfare from 

adverse effects associated jointly with oxides of nitrogen and sulfur. Although EPA has 

historically adopted separate secondary standards for oxides of nitrogen and oxides of sulfur, 

EPA is conducting a joint review of these standards because oxides of nitrogen and sulfur, and 

their associated transformation products are linked from an atmospheric chemistry perspective, 

as well as from an environmental effects perspective. The National Research Council (NRC) has 

recommended that EPA consider multiple pollutants, as appropriate, in forming the scientific 

basis for the NAAQS (NRC, 2004). As discussed in the ISA and REA, there is a strong basis for 

considering these pollutants together, building upon EPA’s past recognition of the interactions of 

these pollutants and on the growing body of scientific information that is now available related to 

these interactions and associated ecological effects.

In defining the scope of this review, it must be considered that EPA has set secondary

standards for two other criteria pollutants related to oxides of nitrogen and sulfur: ozone and 

particulate matter (PM). Oxides of nitrogen are precursors to the formation of ozone in the 

atmosphere, and under certain conditions, can combine with atmospheric ammonia to form 
Page 21 of 238

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 ozonerelated damage to vegetation and PM-related visibility impairment. Protection against those 

effects is provided by the ozone and fine PM secondary standards. This review focuses on 

evaluation of the protection provided by secondary standards for oxides of nitrogen and sulfur 

for two general types of effects: (1) direct effects on vegetation associated with exposure to 

gaseous oxides of nitrogen and sulfur in the ambient air, which are the effects that the current 

NO2 and SO2 secondary standards protect against; and (2) effects associated with the deposition 

of oxides of nitrogen and sulfur to sensitive aquatic and terrestrial ecosystems, including 

deposition in the form of particulate nitrate and particulate sulfate.

The ISA focuses on the ecological effects associated with deposition of ambient oxides of 

nitrogen and sulfur to natural sensitive ecosystems, as distinguished from commercially managed 

forests and agricultural lands. This focus reflects the fact that the majority of the scientific 

evidence regarding acidification and nutrient enrichment is based on studies in unmanaged 

ecosystems. Non-managed terrestrial ecosystems tend to have a higher fraction of nitrogen

deposition resulting from atmospheric nitrogen (US EPA, 2008, section 3.3.2.5). In addition, the 

ISA notes that agricultural and commercial forest lands are routinely fertilized with amounts of 

nitrogen that exceed air pollutant inputs even in the most polluted areas (US EPA, 2008, section 

3.3.9). This review recognizes that the effects of nitrogen deposition in managed areas are 

viewed differently from a public welfare perspective than are the effects of nitrogen deposition in 

natural, unmanaged ecosystems, largely due to the more homogeneous, controlled nature of 
Page 22 of 238

species composition and development in managed ecosystems and the potential for benefits of 

increased productivity in those ecosystems.

In focusing on natural sensitive ecosystems, the PA primarily considers the effects of 

ambient oxides of nitrogen and sulfur via deposition on multiple ecological receptors. The ISA 

highlights effects including those associated with acidification and nitrogen nutrient enrichment. 

With a focus on these deposition-related effects, EPA’s objective is to develop a framework for 

oxides of nitrogen and sulfur standards that incorporates ecologically relevant factors and that 

recognizes the interactions between the two pollutants as they deposit to sensitive ecosystems. 

The overarching policy objective is to develop a secondary standard(s) based on the ecological 

criteria described in the ISA and the results of the assessments in the REA, and consistent with 

the requirement of the CAA to set secondary standards that are requisite to protect the public 

welfare from any known or anticipated adverse effects associated with the presence of these air 

pollutants in the ambient air. Consistent with the CAA, this policy objective includes 

consideration of “variable factors . . . which of themselves or in combination with other factors 

may alter the effects on public welfare” of the criteria air pollutants included in this review.

In addition, we have chosen to focus on the effects of ambient oxides of nitrogen and 

sulfur on ecological impacts on sensitive aquatic ecosystems associated with acidifying 

deposition of nitrogen and sulfur, which is a transformation product of ambient oxides of 

nitrogen and sulfur. Based on the information in the ISA, the assessments presented in the REA, 

and advice from CASAC on earlier drafts of this PA (Russell and Samet, 2010a, 2010b), and as 

discussed in detail in the PA, we have the greatest confidence in the causal linkages between 

oxides of nitrogen and sulfur and aquatic acidification effects relative to other deposition-related 

effects, including terrestrial acidification and aquatic and terrestrial nutrient enrichment.
Page 23 of 238

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 PA, upon which staff 

conclusions on revisions to the secondary oxides of nitrogen and oxides of sulfur standards are 

based; (2) CASAC advice and recommendations, as reflected in discussions of drafts of the ISA, 
Page 24 of 238

REA, and PA at public meetings, in separate written comments, and in CASAC’s letters to the 

Administrator; and (3) public comments received during the development of these documents, 

either in connection with CASAC meetings or separately. 

In developing this rationale, EPA has drawn upon an integrative synthesis of the entire 

body of evidence, published through early 2008, on ecological effects associated with the 

deposition of oxides of nitrogen and oxides of sulfur in the ambient air (US EPA, 2008). As 

discussed below in section II.A, this body of evidence addresses a broad range of ecological 

endpoints associated with ambient levels of oxides of nitrogen and oxides of sulfur. In 

considering this evidence, EPA focuses on those ecological endpoints, such as aquatic 

acidification, for which the ISA judges associations with oxides of nitrogen and oxides of sulfur

to be causal, likely causal, or for which the evidence is suggestive that oxides of nitrogen and/or 

sulfur contribute to the reported effects. The categories of causality determinations have been 

developed in the ISA (US EPA, 2008) and are discussed in Section 1.6 of the ISA.

Crucial to this review is the development of a form for an ecologically relevant standard 

that reflects both the geographically variable and deposition-dependent nature of the effects. 

The atmospheric levels of oxides of nitrogen and sulfur that afford a particular level of 

ecosystem protection are those levels that result in an amount of deposition that is less than the 

amount of deposition that a given ecosystem can accept without defined levels of degradation.

Drawing from the framework developed in the REA, the framework we used to structure 

an ecologically meaningful secondary standard in the PA and to further develop the indicator, 

form, level, and averaging time of such a standard in section III of this proposal is depicted 

below and highlights the three key linkages that need to be considered in developing an 

ecologically relevant standard.
Figure II-1. Simplified conceptual design of the form of an aquatic acidification standard 

for oxides of nitrogen and sulfur.

The following discussion

staff’s evaluation of the policy implications of the scientific assess

presented and assessed in the ISA and the results of quantitative assessments based on that 

information presented and assessed in the REA.

conclusions and the development of policy option

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 

that characterizes emissions, air quality, deposition and water quality.

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 inclu

modeling and water quality data sources. 

Section II.A presents a discussion 

sulfur in the ambient air. The discussion

considered, including direct effects of gaseous oxides of nitrogen and 

effects related to acidification and nutrient enrichment

damage, climate-related effects and mercury methylation

Page 25 of 238

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 PA. The 

staff’s evaluation of the policy implications of the scientific assessment of the evidence 

presented and assessed in the ISA and the results of quantitative assessments based on that 

information presented and assessed in the REA. Taken together, this information informs staff 

conclusions and the development of policy options in the PA for consideration in addressing 

public and welfare effects associated with the presence of oxides of nitrogen and oxides of sulfur 

in the ambient air. Of particular note, chapter 2 of the PA presents information not repeated here 

rizes 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. 

discussion of the effects associated with oxides of nitrogen and 

discussion is organized around the types of effects being 

direct effects of gaseous oxides of nitrogen and sulfur, deposition

acidification and nutrient enrichment, and other effects such as materials 

and mercury methylation.

Simplified conceptual design of the form of an aquatic acidification standard 

The PA includes 

ment of the evidence 

presented and assessed in the ISA and the results of quantitative assessments based on that 

Taken together, this information informs staff 

for consideration in addressing 

public and welfare effects associated with the presence of oxides of nitrogen and oxides of sulfur 

presents information not repeated here 

It includes discussions of 

the sources of nitrogen and sulfur in the atmosphere as well as current ambient air quality 

des ecological 

associated with oxides of nitrogen and 

is organized around the types of effects being 

, deposition-related 

other effects such as materials 
Page 26 of 238

Section II.B presents a summary and discussion of the risk and exposure assessment 

performed for each of the four major effects categories. The REA uses case studies representing 

the broad geographic variability of the impacts from oxides of nitrogen and sulfur to conclude 

that there are ongoing adverse effects in many ecosystems from deposition of oxides of nitrogen 

and sulfur and that under current emissions scenarios these effects are likely to continue.

Section II.C presents a discussion of adversity linking ecological effects to measures that 

can be used to characterize the extent to which such effects are reasonably considered to be 

adverse to public welfare. This involves consideration of how to characterize adversity from a 

public welfare perspective. In so doing, consideration is given to the concept of ecosystem 

services, the evidence of effects on ecosystem services, and how ecosystem services can be 

linked to ecological indicators.

Section II.D presents an assessment of the adequacy of the current oxides of nitrogen and 

oxides of sulfur secondary standards. Consideration is given to the adequacy of protection 

afforded by the current standards for both direct and deposition-related effects, as well as to the 

appropriateness of the fundamental structure and the basic elements of the current standards for 

providing protection from deposition-related effects. Considerations as to the extent to which 

deposition-related effects that could reasonably be judged to be adverse to public welfare are 

occurring under current conditions which are allowed by the current standards is also considered. 

Discussion of the structures and basic elements of the current NO2 and SO2 secondary standards 

and whether they are adequate to protect against such effects is presented.

A. Ecological Effects

This section discusses the known or anticipated ecological effects associated with oxides 

of nitrogen and sulfur, including the direct effects of gas-phase exposure to oxides of nitrogen 
Page 27 of 238

and sulfur (section II.A.1) and effects associated with deposition-related exposure (sections 

II.A.2 and 3). Section II.A. 2 addresses effects related to acidification of aquatic and terrestrial 

ecosystems and section II A.3 addresses effects related to nutrient enrichment of aquatic and 

terrestrial ecosystems. These sections also address questions about the nature and magnitude of 

ecosystem responses to reactive nitrogen and sulfur deposition, including responses related to 

acidification, nutrient depletion, and, in Section II.A 4 the mobilization of toxic metals in 

sensitive aquatic and terrestrial ecosystems. The uncertainties and limitations associated with the 

evidence of such effects are also discussed throughout this section. 

1. Effects Associated with Gas-Phase Oxides of Nitrogen and Sulfur

Ecological effects on vegetation as discussed in earlier reviews as well as the ISA can be 

attributed to gas-phase oxides of nitrogen and sulfur. Acute and chronic exposures to gaseous 

pollutants such as SO2, NO2, nitric oxide (NO), nitric acid (HNO3) and peroxyacetyl nitrite 

(PAN) are associated with negative impacts to vegetation. The current secondary NAAQS were 

set to protect against direct damage to vegetation by exposure to gas-phase oxides of nitrogen 

and sulfur, such as foliar injury, decreased photosynthesis, and decreased growth. The following 

summary is a concise overview of the known or anticipated effects to vegetation caused by gas 

phase nitrogen and sulfur. Most phototoxic effects associated with gas phase oxides of nitrogen

and sulfur occur at levels well above ambient concentrations observed in the U.S. (US EPA, 

2008, section 3.4.2.4).

a. Nature of ecosystem responses to gas-phase nitrogen and sulfur

The 2008 ISA found that gas phase nitrogen and sulfur are associated with direct 

phytotoxic effects (US EPA, 2008, section 4.4). The evidence is sufficient to infer a causal 

relationship between exposure to SO2 and injury to vegetation (US EPA, 2008, section 4.4.1 and 
Page 28 of 238

3.4.2.1). Acute foliar injury to vegetation from SO2 may occur at levels above the current 

secondary standard (3-h average of 0.50 ppm). Effects on growth, reduced photosynthesis and 

decreased yield of vegetation are also associated with increased SO2 exposure concentration and 

time of exposure.

The evidence is sufficient to infer a causal relationship between exposure to NO, NO2

and PAN and injury to vegetation (US EPA, 2008, section 4.4.2 and 3.4.2.2). At sufficient 

concentrations, NO, NO2 and PAN can decrease photosynthesis and induce visible foliar injury

to plants. Evidence is also sufficient to infer a causal relationship between exposure to HNO3

and changes to vegetation (US EPA, 2008, section 4.4.3 and 3.4.2.3). Phytotoxic effects of this 

pollutant include damage to the leaf cuticle in vascular plants and disappearance of some 

sensitive lichen species. 

b. Magnitude of ecosystem response to gas-phase nitrogen and sulfur

Vegetation in ecosystems near sources of gaseous oxides of nitrogen and sulfur or where 

SO2, NO, NO2, PAN and HNO3 are most concentrated are more likely to be impacted by these 

pollutants. Uptake of these pollutants in a plant canopy is a complex process involving 

adsorption to surfaces (leaves, stems and soil) and absorption into leaves (US EPA, 2008, section

3.4.2). The functional relationship between ambient concentrations of gas phase oxides of 

nitrogen and sulfur and specific plant response are impacted by internal factors such as rate of 

stomatal conductance and plant detoxification mechanisms, and external factors including plant 

water status, light, temperature, humidity, and pollutant exposure regime (US EPA, 2008, section

3.4.2).

Entry of gases into a leaf is dependent upon physical and chemical processes of gas phase 

as well as to stomatal aperture. The aperture of the stomata is controlled largely by the 
Page 29 of 238

prevailing environmental conditions, such as water availability, humidity, temperature, and light 

intensity. When the stomata are closed, resistance to gas uptake is high and the plant has a very

low degree of susceptibility to injury. Mosses and lichens do not have a protective cuticle barrier 

to gaseous pollutants or stomata and are generally more sensitive to gaseous sulfur and nitrogen 

than vascular plants (US EPA, 2008, section 3.4.2). 

The appearance of foliar injury can vary significantly across species and growth 

conditions affecting stomatal conductance in vascular plants (US EPA, 2009, section 6.4.1). For 

example, damage to lichens from SO2 exposure includes decreased photosynthesis and 

respiration, damage to the algal component of the lichen, leakage of electrolytes, inhibition of 

nitrogen fixation, decreased potassium (K+) absorption, and structural changes.

The phytotoxic effects of gas phase oxides of nitrogen and sulfur are dependent on the 

exposure concentration and duration and species sensitivity to these pollutants. Effects to 

vegetation associated with oxides of nitrogen and sulfur are therefore variable across the U.S. 

and tend to be higher near sources of photochemical smog. For example, SO2 is considered to be 

the primary factor contributing to the death of lichens in many urban and industrial areas. 

The ISA states there is very limited new research on phytotoxic effects of NO, NO2, PAN 

and HNO3 at concentrations currently observed in the U.S. with the exception of some lichen 

species (US EPA, 2008, section 4.4). Past and current HNO3 concentrations may be contributing 

to the decline in lichen species in the Los Angeles basin. Most phytotoxic effects associated 

with gas phase oxides of nitrogen and sulfur occur at levels well above ambient concentrations 

observed in the U.S. (US EPA, 2008, section 3.4.2.4).

2. Acidification Effects Associated with Deposition of Oxides of Nitrogen and Sulfur
Page 30 of 238

Sulfur oxides and nitrogen oxides in the atmosphere undergo a complex mix of reactions 

in gaseous, liquid, and solid phases to form various acidic compounds. These acidic compounds 

are removed from the atmosphere through deposition: either wet (e.g., rain, snow), fog or cloud, 

or dry (e.g., gases, particles). Deposition of these acidic compounds to ecosystems can lead to 

effects on ecosystem structure and function. Following deposition, these compounds can, in 

some instances, unless retained by soil or biota, leach out of the soils in the form of sulfate 

(SO4

2-

) and nitrate (NO3

-

), leading to the acidification of surface waters. The effects on 

ecosystems depend on the magnitude and rate of deposition, as well as a host of biogeochemical 

processes occurring in the soils and water bodies (US EPA, 2009, section 2.1). The chemical 

forms of nitrogen that may contribute to acidifying deposition include both oxidized and reduced 

chemical species, including reduced forms of nitrogen (NHx).

When sulfur or nitrogen leaches from soils to surface waters in the form of SO4

2-

or NO3

-

, 

an equivalent amount of positive cations, or countercharge, is also transported. This maintains 

electroneutrality. If the countercharge is provided by base cations, such as calcium (Ca2+), 

magnesium (Mg2+), sodium (Na+), or K+, rather than hydrogen (H+) and dissolved inorganic 

aluminum, the acidity of the soil water is neutralized, but the base saturation of the soil 

decreases. Continued SO4

2-

or NO3

-

leaching can deplete the available base cation pool in soil. 

As the base cations are removed, continued deposition and leaching of SO4

2-

and/or NO3

-

(with 

H

+

and 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 
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weathering rates, bedrock composition, vegetation and microbial processes, physical and 

chemical characteristics of soils and hydrologic flowpaths (US EPA, 2009, section 2.1). Some of 

these factors such as vegetation and soil depth are highly variable over small spatial scales such 

as meters, but can be aggregated to evaluate patterns over larger spatial scales. Acidifying 

deposition of oxides of nitrogen and sulfur and the chemical and biological responses associated 

with these inputs vary temporally. Chronic or long-term deposition processes in the time scale of 

years to decades result in increases in inputs of nitrogen and sulfur to ecosystems and the 

associated ecological effects. Episodic or short term (i.e., hours or days) deposition refers to 

events in which the level of the acid neutralizing capacity (ANC) of a lake or stream is 

temporarily lowered. In aquatic ecosystems, short-term (i.e., hours or days) episodic changes in 

water chemistry can have significant biological effects. Episodic acidification refers to 

conditions during precipitation or snowmelt events when proportionately more drainage water is 

routed through upper soil horizons that tend to provide less acid neutralizing than is passing 

through deeper soil horizons (US EPA, 2009, section 4.2). In addition, the accumulated sulfate 

and nitrate in snow packs can provide a surge of acidic inputs. Some streams and lakes may

have chronic or base flow chemistry that is suitable for aquatic biota, but may be subject to 

occasional acidic episodes with deleterious consequences to sensitive biota.

The following summary is a concise overview of the known or anticipated effects caused 

by acidification to ecosystems within the U.S. Acidification affects both terrestrial and 

freshwater aquatic ecosystems. 

a. Nature of acidification-related ecosystem responses

The ISA concluded that deposition of oxides of nitrogen and sulfur and NHx leads to the 

varying degrees of acidification of ecosystems (US EPA, 2008). In the process of acidification, 
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biogeochemical components of terrestrial and freshwater aquatic ecosystems are altered in a way

that leads to effects on biological organisms. Deposition to terrestrial ecosystems often moves 

through the soil and eventually leaches into adjacent water bodies.

i. Aquatic ecosystems

The scientific evidence is sufficient to infer a causal relationship between acidifying 

deposition and effects on biogeochemistry and biota in aquatic ecosystems (US EPA, 2008, 

section 4.2.2). The strongest evidence comes from studies of surface water chemistry in which 

acidic deposition is observed to alter sulfate and nitrate concentrations in surface waters, the sum 

of base cations, ANC, dissolved inorganic aluminum and pH (US EPA, 2008, section 3.2.3.2). 

The ANC is a key indicator of acidification with relevance to both terrestrial and aquatic 

ecosystems. The ANC is useful because it integrates the overall acid-base status of a lake or 

stream and reflects how aquatic ecosystems respond to acidic deposition over time. There is also 

a relationship between ANC and the surface water constituents that directly contribute to or 

ameliorate acidity-related stress, in particular, concentrations of hydrogen ion (as pH), Ca2+ and 

aluminum (Al). Moreover, low pH surface waters leach aluminum from soils, which is quite 

lethal to fish and other aquatic organisms. In aquatic systems, there is a direct relationship 

between ANC and fish and phyto-zooplankton diversity and abundance. 

Low ANC coincides with effects on aquatic systems (e.g., individual species fitness loss

or death, reduced species richness, altered community structure). At the community level,

species richness is positively correlated with pH and ANC because energy cost in maintaining 

physiological homeostasis, growth, and reproduction is high at low ANC levels. For example, 

there is a logistic relationship between fish species richness and ANC class for Adirondack Case 

Study Area lakes that indicates the probability of occurrence of an organism for a given value of 
Page 33 of 238

ANC. Biota are generally not harmed when ANC values are >100 microequivalents per liter

(μeq/L). The number of fish species also peaks at ANC values >100 μeq/L. Below 100 μeq/L 

ANC, fish fitness and community diversity begin to decline (US EPA, section 4.2). Specifically 

at ANC levels between 100 and 50 μeq/L, the fitness of sensitive species (e.g., brook trout, 

zooplankton) begins to decline. 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 increased inorganic Al concentrations (US EPA, 2011, Figures 3-1 and 3-2). 

While ANC level does not cause direct biological harm it is a good overall indicator of the risk 

of acidification (US EPA, 2011, section 3.1.3).

There are clear associations between ANC, pH and aquatic species mortality and health 

which are summarized in section 3.1.1 of the PA. Significant harm to sensitive aquatic species 

has been observed at pH levels below 6. Normal stream pH levels with little to no toxicity range 

from 6 to 7 (MacAvoy et al, 1995). Baker et al (1990) observed that “lakes with pH less than 

approximately 6.0 contain significantly fewer species than lakes with pH levels above 6.0.” As 

noted in Chapter 3, typically at pH <4.5 and an ANC <0 μeq/L, complete to near-complete loss 

of many taxa of organisms occur, including fish and aquatic insect populations, whereas other 
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taxa are reduced to only acidophilic species. Acid Neutralizing Capacity is a measure of how 

much acid can be neutralized in a specific surface water system. An ANC value of 0 or below 

means that surface waters have no ability to neutralize any additional acid inputs.

Additional evidence can help refine the understanding of effects occurring at pH levels 

between 4.5 and 6. When pH levels are below 5.6, relatively lower trout survival rates were 

observed in the 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. 
Page 35 of 238

The highest pH level for any of the studies reported in the ISA is 6.0, suggesting that pH 

above 6.0 is protective against mortality effects for most species. Most thresholds are in the 

range of pH of 5.0 to 6.0, which suggests that a target pH should be no lower than 5.0. 

Protection against mortality in some recreationally important species such as lake trout (pH 

threshold of 5.6) and crappie (pH threshold of 5.5), combined with the evidence of effects on 

larval and embryo survival suggests that pH levels greater than 5.5 should be targeted to provide 

protection against mortality effects throughout the life stages of fish.

Non-lethal effects have been observed at pH levels as high as 6. A study in the 

Shenandoah National Park found that the condition factor, a measure of fish health expressed as 

fish weight/length multiplied by a scaling constant, is positively correlated with stream pH 

levels, and that the condition factor is reduced in streams with a pH of 6.0 (US EPA, 2008, 

section 3.2.3.3).

Biodiversity is another indicator of aquatic ecosystem health. A key study in the 

Adirondacks found that lakes with a pH of 6.0 had only half the potential species of fish (27 of 

53 potential species). There is often a positive relationship between pH and number of fish 

species, at least for pH values between about 5.0 and 6.5, or ANC values between about 0 to 

100 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 
Page 36 of 238

animal species in various ecosystems. These fish may also serve as a source of food and 

recreation for humans. Acidification of ecosystems has been shown to disrupt food web 

dynamics causing alteration to the diet, breeding distribution, and reproduction of certain species 

of birds (US EPA, 2008, section 4.2.2.2. and Table 3-9). For example, breeding distribution of 

the common goldeneye (Bucephala clangula), an insectivorous duck, may be affected by 

changes in acidifying deposition. Similarly, decreases in prey diversity and quantity have been 

observed to create feeding problems for nesting pairs of loons on low-pH lakes in the 

Adirondacks. 

ii. Terrestrial ecosystems

In terrestrial ecosystems, the evidence is sufficient to infer a causal relationship between 

acidifying deposition and changes in biogeochemistry (US EPA, 2008, section 4.2.1.1). The 

strongest evidence comes from studies of forested ecosystems, with supportive information on 

other plant taxa, including shrubs and lichens (US EPA, 2008, section 3.2.2.1.). Three useful 

indicators of chemical changes and acidification effects on terrestrial ecosystems, showing 

consistency and coherence among multiple studies are: soil base saturation, Al concentrations in 

soil water, and soil carbon to nitrogen (C:N) ratio (US EPA, 2008, section 3.2.2.2). 

As discussed in the ISA and REA, in soils with base saturation less than about 15 to 20

percent, exchange chemistry is dominated by Al. Under these conditions, responses to inputs of 

sulfuric acid and HNO3 largely involve the release and mobilization of dissolved inorganic Al. 

The effect can be neutralized by weathering from geologic parent material or base cation 

exchange. The 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 
Page 37 of 238

acidification and cation leaching have been consistently shown to occur only in soils with a C:N 

ratio below about 20 to 25.

Soil acidification caused by acidic deposition has been shown to cause decreased growth 

and increased susceptibility to disease and injury in sensitive tree species. Red spruce (Picea 

rubens) dieback or decline has been observed across high elevation areas in the Adirondack, 

Green and White mountains. The frequency of freezing injury to red spruce needles has 

increased over the past 40 years, a period that coincided with increased emissions of sulfur and 

nitrogen oxides and increased acidifying deposition. Acidifying deposition can contribute to 

dieback in sugar maple (Acer saccharum) through depletion of cations from soil with low levels 

of available Ca. Grasslands are likely less sensitive to acidification than forests due to grassland 

soils being generally rich in base cations.

iii. Ecosystem sensitivity 

The intersection between current deposition loading, historic loading and sensitivity

defines the ecological vulnerability to the effects of acidification. Freshwater aquatic and some 

terrestrial ecosystems, notably forests, are the ecosystem types which are most sensitive to 

acidification. The ISA reports that the principal factor governing the sensitivity of terrestrial and 

aquatic ecosystems to acidification from sulfur and nitrogen deposition is geology (particularly

surficial geology). Geologic formations having low base cation supply generally underlie the 

watersheds of acid-sensitive lakes and streams. Other factors that contribute to the sensitivity of 

soils and surface waters to acidifying deposition include topography, soil chemistry, land use, 

and hydrologic flowpaths. Episodic and chronic acidification tends to occur in areas that have 

base-poor bedrock, high relief, and shallow soils (US EPA, 2008, section 3.2.4.1).
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b. 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.

i. Aquatic acidification

Freshwater ecosystem surveys and monitoring in the eastern U.S. have been conducted 

by many programs since the mid-1980s, including EPA’s Environmental Monitoring and 

Assessment Program (EMAP), National Surface Water Survey (NSWS), Temporally Integrated 

Monitoring of Ecosystems (TIME), and Long-term Monitoring (LTM) programs. Based on 

analyses of surface water data from these programs, New England, the Adirondack Mountains, 

the Appalachian Mountains (northern Appalachian Plateau and Ridge/Blue Ridge region) and the 

Upper Midwest contain the most sensitive lakes and streams (i.e., ANC less than about 50 

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 
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(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).

The ANC is the most widely used indicator of acid sensitivity and has been found in 

various studies to be the best single indicator of the biological response and health of aquatic 

communities in acid-sensitive systems (Lien et al., 1992; Sullivan et al., 2006; US EPA, 2008). 

In the REA, surface water trends in SO4

2-

and NO3

-

concentrations and ANC levels were 

analyzed to affirm the understanding that reductions in deposition could influence the risk of 
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acidification. The ANC values have been categorized according to their effects on biota, as 

shown in the table below. Monitoring data from TIME/LTM and EMAP programs were 

assessed for the years 1990 to 2006, and past, present and future water quality levels were 

estimated by both steady-state and dynamic biogeochemical models.
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Table II-1. Ecological effects associated with alternative levels of acid neutralizing capacity 

(ANC). (source: USEPA, Acid Rain Program)

Category Label ANC Levels and Expected Ecological Effects

Acute 

Concern

<0 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), 
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tiger musky (Esox masquinongy X lucius), walleye (Sander vitreus), alewife (Alosa 

pseudoharengus), and kokanee (Oncorhynchus nerka), as well as ecologically important 

minnows that are commonly consumed by sport fish. A survey of 1,469 lakes in the late 1980s 

found 346 lakes to be devoid of fish. Among lakes with fish, there was a relationship between 

the number of fish species and lake pH, ranging from about one species per lake for lakes having 

a pH <4.5 to about six species per lake for lakes having a pH >6.5. In the Adirondacks, a 

positive relationship exists between the pH and ANC in lakes and the number of fish species 

present in those lakes (US EPA, 2008, section 3.2.3.4).

Since the mid-1990s, streams in the Shenandoah Case Study Area have shown slight 

declines in NO3

-

and SO4

2-

concentrations in surface waters. The 2006 concentrations are still 

above pre-acidification (1860) conditions. Model of Acidification of Groundwater in 

Catchments (MAGIC) modeling predicts surface water concentrations of NO3

-

and SO4

2-

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). Fifty-five percent of all 

monitored streams in the Shenandoah Case Study Area have a current risk of Elevated, Severe, 

or Acute. Of the 55 percent, 18 percent are chronically acidic today (US EPA, 2009, section

4.2.4.3).

Based on a deposition scenario for this study area that maintains current emission levels 

from 2020 to 2050, the simulation forecast indicates that a large number of streams would still 

have Elevated to Acute problems with acidity in 2050.

Biological effects of increased acidification documented in the Shenandoah Case Study

Area include a decrease in the condition factor in blacknose dace and a decrease in fish 

biodiversity associated with decreasing stream ANC. On average, the fish species richness is 
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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).

ii. Terrestrial acidification

The ISA identified a variety of indicators that can be used to measure the effects of 

acidification in soils. Most effects of terrestrial acidification are observed in sensitive forest 

ecosystem in the U.S. Tree health has been linked to the availability of base cations (BC) in soil 

(such as Ca2+, Mg2+ and K

+

), as well as soil aluminum (Al) content. Tree species show a range 

of sensitivities to Ca/Al and BC/Al soil molar ratios, therefore these are good chemical indicators 

because they directly relate to the biological effects. Critical BC/Al molar ratios for a large 

variety of tree species ranged from 0.2 to 0.8. This range is similar to critical ratios of Ca/Al. 

Plant toxicity or nutrient antagonism was reported to occur at Ca/Al molar ratios ranging from 

0.2 to 2.5 (US EPA, 2009).

There has been no systematic national survey of terrestrial ecosystems to determine the 

extent and distribution of terrestrial ecosystem sensitivity to the effects of acidifying deposition. 

However, one preliminary national evaluation estimated that ~15 percent of forest ecosystems in 

the U.S. exceed the estimated critical load based on soil ANC leaching for sulfur and nitgogen

deposition by >250 eq/ha/yr (McNulty et al., 2007). Forests of the Adirondack Mountains of 

New York, Green Mountains of Vermont, White Mountains of New Hampshire, the Allegheny 

Plateau of Pennsylvania and high-elevation forest ecosystems in the southern Appalachians are 

the regions most sensitive to terrestrial acidification effects from acidifying deposition (US EPA, 

2008, section 3.2.4.2). While studies show some recovery of surface waters, there are 

widespread measurements of ongoing depletion of exchangeable base cations in forest soils in 
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the northeastern U.S. despite recent decreases in acidifying deposition, indicating a slow 

recovery time.

In the REA, a critical load analysis was performed for sugar maple and red spruce forests 

in the eastern U.S. by using BC/Al ratio in acidified forest soils as an indicator to assess the 

impact of nitrogen and sulfur deposition on tree health. These are the two most commonly

studied tree species in North America for effects of acidification. At a BC/Al ratio of 1.2, red 

spruce growth can be decreased by 20 percent. Sugar maple growth can be decreased by 20

percent at a BC/Al ratio of 0.6 (US EPA, 2009, section 4.4). The REA analysis determined the 

health of at least a portion of the sugar maple and red spruce growing in the U.S. may have been 

compromised with acidifying total nitrogen and sulfur deposition. Specifically, total nitrogen 

and sulfur deposition levels exceeded three selected critical loads for tree growth in 3 percent to 

75 percent of all sugar maple plots across 24 states -- that is, it exceeded the highest (least 

stringent) of the three critical loads in 3 percent of plots, and the lowest (most stringent) in 75

percent of plots. For red spruce, total nitrogen and sulfur deposition levels exceeded three 

selected critical loads in 3 percent to 36 percent of all red spruce plots across eight states (US 

EPA, 2009, section 4.4). 

c. Key uncertainties associated with acidification

There are different levels of uncertainty associated with relationships between deposition, 

ecological effects and ecological indicators. In Chapter 7 of the REA, the case study analyses 

associated with each targeted effect area were synthesized by identifying the strengths, 

limitations, and uncertainties associated with the available data, modeling approach, and 

relationship between the selected ecological indicator and atmospheric deposition as described 

by the ecological effect function (US EPA, 2009, Figure 1-1). A further discussion of 
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uncertainty in aquatic and terrestrial ecosystems is presented below. The key uncertainties were 

characterized as follows to evaluate the strength of the scientific basis for setting a national 

standard to protect against a given effect (US EPA, 2009, section 7):

(1) Data Availability: high, medium or low quality. This criterion is based on the availability

and robustness of data sets, monitoring networks, availability of data that allows for 

extrapolation to larger assessment areas and input parameters for modeling and developing 

the ecological effect function. The scientific basis for the ecological indicator selected is also 

incorporated into this criterion.

(2) Modeling Approach: high, fairly high, intermediate, or low confidence. This value is based 

on the strengths and limitations of the models used in the analysis and how accepted they are 

by the scientific community for their application in this analysis.

(3) Ecological Effect Function: high, fairly high, intermediate or low confidence. This ranking 

is based on how well the ecological effect function describes the relationship between 

atmospheric deposition and the ecological indicator of an effect.

i. Aquatic acidification

The REA concludes that the available data are robust and considered high quality. There 

is high confidence about the use of these data and their value for extrapolating to a larger 

regional population of lakes. The EPA TIME/LTM network represents a source of long-term, 

representative sampling. Data on sulfate concentrations, nitrate concentrations and ANC from 

1990 to 2006 used for this analysis as well as EPA EMAP and Regional Environmental 

Monitoring and Assessment Program (REMAP) surveys, provide considerable data on surface 

water trends. 
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There is fairly high confidence associated with modeling and input parameters. 

Uncertainties in water quality estimates (i.e., ANC) from MAGIC were derived from multiple 

site calibrations. Pre-acidification refers to retrospective modeling to estimate water quality

conditions before man-made contributions of acidifying inputs. The models are evaluated under

current conditions to determine how well they replicate observed ANC values. The 95 percent 

confidence interval for pre-acidification of lakes was an average of 15 μeq/L difference in ANC 

concentrations, or 10 percent, and 8 μeq/L, or 5 percent, for streams (US EPA, 2009, section

7.1.2). The use of the critical load model to estimate aquatic critical loads is limited by the 

uncertainties associated with runoff and surface water measurements and in estimating the 

catchment supply of base cations from the weathering of bedrock and soils (McNulty et al., 

2007).

ii. Terrestrial acidification 

The available data used to quantify the targeted effect of terrestrial acidification are 

robust and considered high quality. The U.S. Forest Service-Kane Experimental Forest and 

significant amounts of research work in the Allegheny Plateau have produced extensive, peerreviewed data sets. 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 
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the REA analysis (US EPA, 2008, section 7.2.2). There is fairly high confidence associated with 

the ecological effect function developed for terrestrial acidification (US EPA, 2009, section

7.2.3).

3. Nutrient Enrichment Effects Associated with Deposition of Oxides of Nitrogen 

The following summary is a concise overview of the known or anticipated effects caused 

by nitrogen nutrient enrichment to ecosystems within the United States. Nutrient-enrichment 

affects terrestrial, freshwater and estuarine ecosystems. Nitrogen deposition is a major source of 

anthropogenic nitrogen. For many terrestrial and freshwater ecosystems other sources of 

nitrogen including fertilizer and waste treatment are greater than deposition. Nitrogen deposition 

often contributes to nitrogen-enrichment effects in estuaries, but does not drive the effects since 

other sources of nitrogen greatly exceed nitrogen deposition. Both oxides of nitrogen and NHx

contribute to nitrogen deposition. For the most part, nitrogen effects on ecosystems do not 

depend on whether the nitrogen is in oxidized or reduced form. Thus, this summary focuses on 

the effects of nitrogen deposition in total. 

a. Nature of nutrient enrichment-related ecosystem responses

The ISA found that deposition of nitrogen, including oxides of nitrogen and NHx, leads to 

the nitrogen enrichment of ecosystems (US EPA 2008). In the process of nitrogen enrichment, 

biogeochemical components of terrestrial and freshwater aquatic ecosystems are altered in a way

that leads to effects on biological organisms. 

i. Aquatic ecosystems

In freshwater ecosystems, the evidence is sufficient to infer a causal relationship between 

nitrogen deposition and the alteration of biogeochemical cycling in freshwater aquatic 

ecosystems (US EPA, 2008, section 3.3.2.3). Nitrogen deposition is the main source of nitrogen
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enrichment to headwater streams, lower order streams and high elevation lakes. The most 

common chemical indicators that were studied included NO3

−

and dissolved inorganic nitrogen 

(DIN) concentration in surface waters as well as the ratio of chlorophyll a to total phosphorus.

Elevated surface water NO3

−

concentrations occur in both the eastern and western U.S. Studies 

report a significant correlation between nitrogen deposition and lake biogeochemistry by

identifying a correlation between wet deposition and DIN and the ratio of chlorophyll a to total

phosphate. Recent evidence provides examples of lakes and streams that are limited by nitrogen

and show signs of eutrophication in response to nitrogen addition.

The evidence is sufficient to infer a causal relationship between nitrogen deposition and 

the alteration of species richness, species composition and biodiversity in freshwater aquatic 

ecosystems (US EPA, 2008, section 3.3.5.3). Increased nitrogen deposition can cause a shift in 

community composition and reduce algal biodiversity, especially in sensitive oligotrophic lakes.

In the ISA, the evidence is sufficient to infer a causal relationship between nitrogen 

deposition and the biogeochemical cycling of nitrogen and carbon in estuaries (US EPA, 2008, 

section 4.3.4.1 and 3.3.2.3). In general, estuaries tend to be nitrogen-limited, and many currently 

receive high levels of nitrogen input from human activities (US EPA, 2009, section 5.1.1). It is 

unknown if atmospheric deposition alone is sufficient to cause eutrophication; however, the 

contribution of atmospheric nitrogen deposition to total nitrogen load is calculated for some 

estuaries and can be >40 percent (US EPA, 2009, section 5.1.1).

The evidence is sufficient to infer a causal relationship between nitrogen deposition and 

the alteration of species richness, species composition and biodiversity in estuarine ecosystems

(US EPA, 2008, section 4.3.4.2 and 3.3.5.4). Atmospheric and non-atmospheric sources of 

nitrogen contribute to increased phytoplankton and algal productivity, leading to eutrophication.
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Shifts in community composition, reduced hypolimnetic dissolved oxygen (DO), decreases in 

biodiversity, and mortality of submerged aquatic vegetation are associated with increased N 

deposition in estuarine systems. 

ii. Terrestrial Ecosystems

The evidence is sufficient to infer a causal relationship between nitrogen deposition and 

the alteration of biogeochemical cycling in terrestrial ecosystems (US EPA, 2008, section 4.3.1.1 

and 3.3.2.1). This is supported by numerous observational, deposition gradient and field addition 

experiments in sensitive ecosystems. The leaching of NO3- in soil drainage waters and the 

export of NO3- in stream water were identified as two of the primary indictors of nitrogen

enrichment. Several nitrogen-addition studies indicate that NO3- leaching is induced by chronic 

additions of nitrogen. Studies identified in the ISA found that surface water NO3-

concentrations exceeded 1 μeq/L in watersheds receiving about 9 to 13 kg N/ha/yr of 

atmospheric nitrogen deposition. Nitrogen deposition disrupts the nutrient balance of 

ecosystems with numerous biogeochemical effects. . The chemical indicators that are typically 

measured include NO3

−

leaching, soil C:N ratio, rates of nitrogen mineralization, nitrification, 

denitrification, foliar nitrogen concentration, and soil water NO3 − and NH4

+

concentrations. 

Note that nitrogen saturation (nitrogen leaching from ecosystems) does not need to occur to 

cause effects. Substantial leaching of NO3− from forest soils to stream water can acidify

downstream waters, leading to effects described in the previous section on aquatic acidification. 

Due to the complexity of interactions between the nitrogen and carbon cycling, the effects of 

nitrogen on carbon budgets (quantified input and output of carbon to the ecosystem) are variable. 

Regional trends in net ecosystem productivity (NEP) of forests (not managed for silviculture) 

have been estimated through models based on gradient studies and meta-analysis. Atmospheric 
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nitrogen deposition has been shown to cause increased litter accumulation and carbon storage in 

above-ground woody biomass. In the West, this has lead to increased susceptibility to more 

severe fires. Less is known regarding the effects of nitrogen deposition on carbon budgets of 

non-forest ecosystems.

The evidence is sufficient to infer a causal relationship between nitrogen deposition on 

the alteration of species richness, species composition and biodiversity in terrestrial ecosystems

(US EPA, 2008, section 4.3.1.2). Some organisms and ecosystems are more sensitive to nitrogen

deposition and effects of nitrogen deposition are not observed in all habitats. The most sensitive 

terrestrial taxa to nitrogen deposition are lichens. Empirical evidence indicates that lichens in the 

U.S. are affected by deposition levels as low as 3 kg N/ha/yr. Alpine ecosystems are also 

sensitive to nitrogen deposition; changes in an individual species (Carex rupestris) were 

estimated to occur at deposition levels near 4 kg N /ha/yr and modeling indicates that deposition 

levels near 10 kg N/ha/yr alter plant community assemblages. In several grassland ecosystems, 

reduced species diversity and an increase in non-native, invasive species are associated with 

nitrogen deposition. 

iii. Ecosystem sensitivity to nutrient enrichment

The numerous ecosystem types that occur across the U.S. have a broad range of 

sensitivity to nitrogen deposition (US EPA, 2008, Table 4-4). Increased deposition to nitrogenlimited ecosystems can lead to production increases that may be either beneficial or adverse 

depending on the system and management goals. 

Organisms in their natural environment are commonly adapted to a specific regime of 

nutrient availability. Change in the availability of one important nutrient, such as nitrogen, may

result in an imbalance in ecological stoichiometry, with effects on ecosystem processes, structure 
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and function. In general, nitrogen deposition to terrestrial ecosystems causes accelerated growth 

rates in some species deemed desirable in commercial forests but may lead to altered competitive 

interactions among species and nutrient imbalances, ultimately affecting biodiversity. The onset 

of these effects occurs with nitrogen deposition levels as low as 3 kg N/ha/yr in sensitive 

terrestrial ecosystems to nitrogen deposition. In aquatic ecosystems, nitrogen that is both 

leached from the soil and directly deposited to the water surface can pollute the surface water. 

This causes alteration of the diatom community at levels as low as 1.5 kg N/ha/yr in sensitive 

freshwater ecosystems. 

The degree of ecosystem effects lies at the intersection of nitrogen loading and nitrogensensitivity. Nitrogen-sensitivity is predominately driven by the degree to which growth is 

limited by nitrogen availability. Grasslands in the western U.S. are typically nitrogen-limited 

ecosystems dominated by a diverse mix of perennial forbs and grass species. A meta-analysis 

discussed in the ISA (US EPA, 2008, section 3.3.3), indicated that nitrogen fertilization 

increased aboveground growth in all non-forest ecosystems except for deserts. In other words, 

almost all terrestrial ecosystems are nitrogen-limited and will be altered by the addition of 

anthropogenic nitrogen. Likewise, a freshwater lake or stream must be nitrogen-limited to be 

sensitive to nitrogen-mediated eutrophication. There are many examples of fresh waters that are 

nitrogen-limited or nitrogen and phosphorous (P) co-limited (US EPA, 2008, section 3.3.3.2). A 

large dataset meta-analysis discussed in the ISA (US EPA, 2008, section 3.3.3.2), found that 

nitrogen-limitation occurred as frequently as phosphorous-limitation in freshwater ecosystems. 

Additional factors that govern the sensitivity of ecosystems to nutrient enrichment from nitrogen

deposition include rates and form of nitrogen deposition, elevation, climate, species composition, 

plant growth rate, length of growing season, and soil nitrogen retention capacity (US EPA, 2008, 
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section 4.3). Less is known about the extent and distribution of the terrestrial ecosystems in the 

U.S. that are most sensitive to the effects of nutrient enrichment from atmospheric nirogen

deposition compared to acidification.

Because the productivity of estuarine and near shore marine ecosystems is generally 

limited by the availability of nitrogen, they are susceptible to the eutrophication effect of 

nitrogen deposition (US EPA, 2008, section 4.3.4.1). A recent national assessment of eutrophic 

conditions in estuaries found the most eutrophic estuaries were generally those that had large 

watershed-to-estuarine surface area, high human population density, high rainfall and runoff, low 

dilution and low flushing rates. In the REA, the National Oceanic and Atmospheric 

Administration’s (NOAA) National Estuarine Eutrophication Assessment (NEEA) assessment 

tool, Assessment of Estuarine Tropic Status (ASSETS) categorical Eutrophication Index (EI) 

was used to evaluate eutrophication due to atmospheric loading of nitrogen. The ASSETS EI is 

an estimation of the likelihood that an estuary is experiencing eutrophication or will experience 

eutrophication based on five ecological indicators: chlorophyll a, macroalgae, dissolved oxygen, 

nuisance/toxic algal blooms and submerged aquatic vegetation (SAV). 

In the REA, two regions were selected for case study analysis using ASSETS EI, the 

Chesapeake Bay and Pamlico Sound. Both regions received an ASSETS EI rating of Bad

indicating that the estuary had moderate to high pressure due to overall human influence and a 

moderate high to high eutrophic condition (US EPA, 2009, sections 5.2.4.1 and 5.2.4.2). These 

results were then considered with SPAtially Referenced Regression on Watershed Attributes

(SPARROW) modeling to develop a response curve to examine the role of atmospheric nitrogen 

deposition in achieving a desired decrease in load. To change the Neuse River Estuary’s EI 

score from Bad to Poor not only must 100 percent of the total atmospheric nitrogen deposition be 
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eliminated, but considerably more nitrogen from other sources as well must be controlled (US 

EPA, 2009, section 5.2.7.2). In the Potomac River estuary, a 78 percent decrease of total 

nitrogen could move the EI score from Bad to Poor (US EPA, 2009, section 5.2.7.1). The results 

of this analysis indicated decreases in atmospheric deposition alone could not eliminate coastal 

eutrophication problems due to multiple non-atmospheric nitrogen inputs (US EPA, 2009, 

section 7.3.3). However, the somewhat arbitrary discreteness of the EI scale can mask the 

benefits of decreases in nitrogen between categories.

In general, estuaries tend to be nitrogen-limited, and many currently receive high levels 

of nitrogen input from human activities to cause eutrophication. As reported in the ISA (US 

EPA, 2008, section 3.2.2.2), atmospheric nitrogen loads to estuaries in the U.S. are estimated to 

range from 2 to 8 percent for Guadalupe Bay, Texas on the lowest end to as high as 72 percent

for St. Catherines-Sapelo estuary, Georgia. The Chesapeake Bay is an example of a large, wellstudied and severely eutrophic estuary that is calculated to receive as much as 30 percent of its 

total nitrogen load from the atmosphere.

b. Magnitude of ecosystem responses

i. Aquatic ecosystems

The magnitude of ecosystem response may be thought of on two time scales, current 

conditions and how ecosystems have been altered since the onset of anthropogenic nitrogen

deposition. As noted previously, studies found that nitrogen-limitation occurs as frequently as 

phosphorous-limitation in freshwater ecosystems (US EPA, 2008, section 3.3.3.2). Recently, a 

comprehensive study of available data from the northern hemisphere surveys of lakes along 

gradients of nitrogen deposition show increased inorganic nitrogen concentration and 

productivity to be correlated with atmospheric nitrogen deposition. The results are unequivocal 
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evidence of nitrogen limitation in lakes with low ambient inputs of nitrogen, and increased 

nitrogen concentrations in lakes receiving nitrogen solely from atmospheric nitrogen deposition. 

It has been suggested that most lakes in the northern hemisphere may have originally been 

nitrogen-limited, and that atmospheric nitrogen deposition has changed the balance of nitrogen

and phosphorous in lakes.

Available data suggest that the increases in total nitrogen deposition do not have to be 

large to elicit an ecological effect. For example, a hindcasting exercise determined that the 

change in Rocky Mountain National Park lake algae that occurred between 1850 and 1964 was 

associated with an increase in wet nitrogen deposition that was only about 1.5 kg N/ha. Similar 

changes inferred from lake sediment cores of the Beartooth Mountains of Wyoming also 

occurred at about 1.5 kg N/ha deposition. Pre-industrial inorganic nitrogen deposition is 

estimated to have been only 0.1 to 0.7 kg N/ha based on measurements from remote parts of the 

world. In the western U.S., pre-industrial, or background, inorganic nitrogen deposition was 

estimated by to range from 0.4 to 0.7 kg N/ha/yr.

Eutrophication effects from nitrogen deposition are most likely to be manifested in 

undisturbed, low nutrient surface waters such as those found in the higher elevation areas of the

western U.S. The most severe eutrophication from nitrogen deposition effects is expected 

downwind of major urban and agricultural centers. High concentrations of lake or streamwater 

NO3

−

, indicative of ecosystem saturation, have been found at a variety of locations throughout 

the U.S., including the San Bernardino and San Gabriel Mountains within the Los Angeles Air 

Basin, the Front Range of Colorado, the Allegheny mountains of West Virginia, the Catskill 

Mountains of New York, the Adirondack Mountains of New York, and the Great Smoky 

Mountains in Tennessee (US EPA, 2008, section 3.3.8).
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In contrast to terrestrial and freshwater systems, atmospheric nitrogen load to estuaries 

contributes to the total load but does not necessarily drive the effects since other combined 

sources of nitrogen often greatly exceed nitrogen deposition. In estuaries, nitrogen-loading from 

multiple anthropogenic and non-anthropogenic pathways leads to water quality deterioration, 

resulting in numerous effects including hypoxic zones, species mortality, changes in community 

composition and harmful algal blooms that are indicative of eutrophication. The following 

summary is a concise overview of the known or anticipated effects of nitrogen enrichment on 

estuaries within the U.S.

There is a scientific consensus (US EPA, 2008, section 4.3.4) that nitrogen-driven 

eutrophication in shallow estuaries has increased over the past several decades and that the 

environmental degradation of coastal ecosystems due to nitrogen, phosphorus, and other inputs is 

now a widespread occurrence. For example, the frequency of phytoplankton blooms and the 

extent and severity of hypoxia have increased in the Chesapeake Bay and Pamlico estuaries in 

North Carolina and along the continental shelf adjacent to the Mississippi and Atchafalaya 

rivers’ discharges to the Gulf of Mexico. 

A recent national assessment of eutrophic conditions in estuaries found that 65 percent of 

the assessed systems had moderate to high overall eutrophic conditions Most eutrophic

estuaries occurred in the mid-Atlantic region and the estuaries with the lowest degree of 

eutrophication were in the North Atlantic. Other regions had mixtures of low, moderate, and 

high degrees of eutrophication (US EPA, 2008, section 4.3.4.3).

The mid-Atlantic region is the most heavily impacted area in terms of moderate or high 

loss of submerged aquatic vegetation due to eutrophication (US EPA, 2008, section 4.3.4.2). 

Submerged aquatic vegetation is important to the quality of estuarine ecosystem habitats because 
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it provides habitat for a variety of aquatic organisms, absorbs excess nutrients, and traps 

sediments (US EPA, 2008, section 4.3.4.2). It is partly because many estuaries and near-coastal 

marine waters are degraded by nutrient enrichment that they are highly sensitive to potential 

negative impacts from nitrogen addition from atmospheric deposition.

ii. Terrestrial ecosystems

Little is known about the full extent and distribution of the terrestrial ecosystems in the 

U.S. that are most sensitive to impacts caused by nutrient enrichment from atmospheric nitrogen

deposition. As previously stated, most terrestrial ecosystems are nitrogen-limited, therefore they

are sensitive to perturbation caused by nitrogen additions (US EPA, 2008, section 4.3.1). Effects 

are most likely to occur where areas of relatively high atmospheric N deposition intersect with 

nitrogen-limited plant communities. The alpine ecosystems of the Colorado Front Range, 

chaparral watersheds of the Sierra Nevada, lichen and vascular plant communities in the San 

Bernardino Mountains and the Pacific Northwest, and the southern California coastal sage scrub 

(CSS) community are among the most sensitive terrestrial ecosystems. There is growing 

evidence (US EPA, 2008, section 4.3.1.2) that existing grassland ecosystems in the western U.S. 

are being altered by elevated levels of N inputs, including inputs from atmospheric deposition.

In the eastern U.S., the degree of nitrogen saturation of the terrestrial ecosystem is often 

assessed in terms of the degree of NO3

−

leaching from watershed soils into ground water or 

surface water. Studies have estimated the number of surface waters at different stages of 

saturation across several regions in the eastern U.S. Of the 85 northeastern watersheds examined 

60 percent were in Stage 1 or Stage 2 of nitrogen saturation on a scale of 0 (background or 

pretreatment) to 3 (visible decline). Of the northeastern sites for which adequate data were 

available for assessment, those in Stage 1 or 2 were most prevalent in the Adirondack and 
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Catskill Mountains. Effects on individual plant species have not been well studied in the U.S. 

More is known about the sensitivity of particular plant communities. Based largely on results 

obtained in more extensive studies conducted in Europe, it is expected that the more sensitive 

terrestrial ecosystems include hardwood forests, alpine meadows, arid and semi-arid lands, and 

grassland ecosystems (US EPA, 2008, section 3.3.5).

The REA used published research results (US EPA, 2009, section 5.3.1 and US EPA, 

2008, Table 4.4) to identify meaningful ecological benchmarks associated with different levels of 

atmospheric nitrogen deposition. These are illustrated in Figure 3-4 of the PA. The sensitive 

areas and ecological indicators identified by the ISA were analyzed further in the REA to create 

a national map that illustrates effects observed from ambient and experimental atmospheric 

nitrogen deposition loads in relation to Community Multi-scale Air Quality (CMAQ) 2002 

modeling results and National Atmospheric Deposition Program (NADP) monitoring data. This 

map, reproduced in Figure 3-5 of the PA, depicts the sites where empirical effects of terrestrial 

nutrient enrichment have been observed and site proximity to elevated atmospheric nitrogen

deposition. 

Based on information in the ISA and initial analysis in the REA, further case study

analyses on terrestrial nutrient enrichment of ecosystems were developed for the CS community

and Mixed Conifer Forest (MCF) (US EPA, 2009). Geographic information systems (GIS) 

analysis supported a qualitative review of past field research to identify ecological benchmarks 

associated with CSS and mycorrhizal communities, as well as MCF nutrient-sensitive acidophyte 

lichen communities, fine-root biomass in Ponderosa pine, and leached nitrate in receiving waters. 

The ecological benchmarks that were identified for the CSS and the MCF communities

are included in the suite of benchmarks identified in the ISA (US EPA, 2008, section 3.3). There 
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are sufficient data to confidently relate the ecological effect to a loading of atmospheric nitrogen.

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

(1) 3.3 kg N/ha/yr – the amount of nitrogen uptake by a vigorous stand of CSS; above this level, 

nitrogen may no longer be limiting

(2) 10 kg N/ha/yr – mycorrhizal community changes

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

(1) 3.1 kg N/ha/yr – shift from sensitive to tolerant lichen species

(2) 5.2 kg N/ha/yr – dominance of the tolerant lichen species

(3) 10.2 kg N/ha/yr – loss of sensitive lichen species

(4) 17 kg N/ha/yr – leaching of nitrate into streams.

These benchmarks, ranging from 3.1 to 17 kg N/ha/yr, were compared to 2002 

CMAQ/NADP data to discern any associations between atmospheric deposition and changing 

communities. Evidence supports the finding that nitrogen alters CSS and MCF communities.

Key findings include the following: 2002 CMAQ/NADP nitrogen deposition data show that the 

3.3 kg N/ha/yr benchmark has been exceeded in more than 93 percent of CSS areas (654,048 ha).

These deposition levels are a driving force in the degradation of CSS communities. Although 

CSS decline has been observed in the absence of fire, the contributions of deposition and fire to 

the CSS decline require further research. The CSS is fragmented into many small parcels, and 

the 2002 CMAQ/NADP 12-km grid data are not fine enough to fully validate the relationship 

between CSS distribution, nitrogen deposition, and fire. The 2002 CMAQ/NADP nitrogen 

deposition data exceeds the 3.1 kg N/ha/yr benchmark in more than 38 percent (1,099,133 ha) of 

MCF areas, and nitrate leaching has been observed in surface waters. Ozone effects confound 
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nitrogen effects on MCF acidophyte lichen, and the interrelationship between fire and nitrogen 

cycling requires additional research.

c. Key uncertainties associated with nutrient enrichment

There are different levels of uncertainty associated with relationships between deposition, 

ecological effects and ecological indicators. The criteria used in the REA to evaluate the degree 

of confidence in the data, modeling and ecological effect function are detailed in chapter 7 of the 

REA. Below is a discussion of uncertainty relating aquatic and terrestrial ecosystems to nutrient 

enrichment effects.

i. Aquatic ecosystems 

The approach for assessing atmospheric contributions to total nitrogen loading in the 

REA was to consider the main-stem river to an estuary (including the estuary) rather than an 

entire estuary system or bay. The biological indicators used in the NOAA ASSETS EI required 

the evaluation of many national databases including the US Geological Survey National Water 

Quality Assessment (NAWQA) files, EPA’s STORage and RETrieval (STORET) database, 

NOAA’s Estuarine Drainage Areas data and EPA’s water quality standards nutrient criteria for 

rivers and lakes (US EPA, 2009, Appendix 6 and Table 1.2.-1). Both the SPARROW modeling 

for nitrogen loads and assessment of estuary conditions under NOAA ASSETS EI, have been 

applied on a national scale. The REA concludes that the available data are medium quality with 

intermediate confidence about the use of these data and their values for extrapolating to a larger 

regional area (US EPA, 2009, section 7.3.1). Intermediate confidence is associated with the 

modeling approach using ASSETS EI and SPARROW. The REA states there is low confidence 

with the ecological effect function due to the results of the analysis which indicated that 
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reductions in atmospheric deposition alone could not solve coastal eutrophication problems due 

to multiple non-atmospheric nitrogen inputs (US EPA, 2009, section 7.3.3).

ii. Terrestrial ecosystems

Ecological thresholds are identified for CSS and MCF areas and these data are considered 

to be of high quality, however, the ability to extrapolate these data to larger regional areas is 

limited (US EPA, 2009, section 7.4.1). No quantitative modeling was conducted or ecological 

effect function developed for terrestrial nutrient enrichment reflecting the uncertainties 

associated with these depositional effects. 

4. Other Ecological Effects 

It is stated in the ISA (US EPA, 2008, section 3.4.1 and 4.5) that mercury is a highly 

neurotoxic contaminant that enters the food web as a methylated compound, methylmercury

(MeHg). Mercury is principally methylated by sulfur-reducing bacteria and can be taken up by

microorganisms, zooplankton and macroinvertebrates. The contaminant is concentrated in higher 

trophic levels, including fish eaten by humans. Experimental evidence has established that only 

inconsequential amounts of MeHg can be produced in the absence of sulfate. Once MdHg is 

present, other variables influence how much accumulates in fish, but elevated mercury levels in 

fish can only occur where substantial amounts of MeHg are present. Current evidence indicates 

that in watersheds where mercury is present, increased oxides of sulfur deposition very likely 

results in additional production of MeHg which leads to greater accumulation of MeHg 

concentrations in fish. With respect to sulfur deposition and mercury methylation, the final ISA 

determined that “[t]he evidence is sufficient to infer a causal relationship between sulfur 

deposition and increased mercury methylation in wetlands and aquatic environments.” 
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The production of meaningful amounts of MeHg requires the presence of SO4

2-

and 

mercury, and where mercury is present, increased availability of SO4

2-

results in increased 

production of MeHg. There is increasing evidence on the relationship between sulfur deposition 

and increased methylation of mercury in aquatic environments; this effect occurs only where 

other factors are present at levels within a range to allow methylation. The production of MeHg 

requires the presence of SO4

2-

and mercury, but the amount of MeHg produced varies with 

oxygen content, temperature, pH, and supply of labile organic carbon (US EPA, 2008, section

3.4). In watersheds where changes in sulfate deposition did not produce an effect, one or several 

of those interacting factors were not in the range required for meaningful methylation to occur 

(US EPA, 2008, section 3.4). Watersheds with conditions known to be conducive to mercury

methylation can be found in the northeastern U.S. and southeastern Canada. 

While the relationship between sulfur and MeHg production was concluded to be causal 

in the ISA, the REA concluded that there was insufficient evidence to quantify the relationship 

between sulfur and MeHg. Therefore only a qualitative assessment was included in chapter 6 of 

the REA. The PA was then unable to make a determination as to the adequacy of the existing 

SO2 standards in protecting against welfare effects associated with increased mercury

methylation.

B. Risk and Exposure Assessment 

The risk and exposure assessment conducted for the current review was developed to 

describe potential risk from current and future deposition of oxides of nitrogen and sulfur to 

sensitive ecosystems. The case study analyses in the REA show that there is confidence that 

known or anticipated adverse ecological effects are occurring under current ambient loadings of 

nitrogen and sulfur in sensitive ecosystems across the U.S. An overview of the material covered 
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in the REA, a summary of the key findings from the air quality analyses, acidification and 

nutrient enrichment case studies, and general conclusions from evaluating additional welfare 

effects, are presented below.

1. Overview of the Risk and Exposure Assessment

The REA evaluates the relationships between atmospheric concentrations, deposition, 

biologically relevant exposures, targeted ecosystem effects, and ecosystem services. To evaluate 

the nature and magnitude of adverse effects associated with deposition, the REA also examines 

various ways to quantify the relationships between air quality indicators, deposition of 

biologically available forms of nitrogen and sulfur, ecologically relevant indicators relating to 

deposition, exposure and effects on sensitive receptors, and related effects resulting in changes in 

ecosystem structure and services. The intent is to determine the exposure metrics that 

incorporate the temporal considerations (i.e., biologically relevant timescales), pathways, and 

ecologically relevant indicators necessary to determine the effects on these ecosystems. To the 

extent feasible, the REA evaluates the overall load to the system for nitrogen and sulfur, as well 

as the variability in ecosystem responses to these pollutants. It also evaluates the contributions 

of atmospherically deposited nitrogen and sulfur individually relative to the combined 

atmospheric loadings of both elements together.. Since oxidized nitrogen is the listed criteria 

pollutant (currently measured by the ambient air quality indicator NO2) for the atmospheric 

contribution to total nitrogen, the REA examines the contribution of nitrogen oxides to total 

reactive nitrogen in the atmosphere, relative to the contributions of reduced forms of nitrogen 

(e.g., ammonia, ammonium), to ultimately assess how a meaningful secondary NAAQS might be 

structured.
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The REA focuses on ecosystem welfare effects that result from the deposition of total 

reactive nitrogen and sulfur. Because ecosystems are diverse in biota, climate, geochemistry, 

and hydrology, response to pollutant exposures can vary greatly between ecosystems. In 

addition, these diverse ecosystems are not distributed evenly across the United States. To target 

nitrogen and sulfur acidification and nitrogen and sulfur enrichment, the REA addresses four 

main targeted ecosystem effects on terrestrial and aquatic systems identified by the ISA (US 

EPA, 2008): aquatic acidification due to nitrogen and sulfur; terrestrial acidification due to 

nitrogen and sulfur; aquatic nutrient enrichment, including eutrophication; and terrestrial nutrient 

enrichment.

In addition to these four targeted ecosystem effects, the REA also qualitatively addresses 

the influence of sulfur oxides deposition on MeHg production; nitrous oxide (N2O) effects on 

climate; nitrogen effects on primary productivity and biogenic greenhouse gas (GHG) fluxes; 

and phytotoxic effects on plants. 

Because the targeted ecosystem effects outlined above are not evenly distributed across 

the U.S., the REA identified case studies for each targeted effects based on ecosystems identified 

as sensitive to nitrogen and/or sulfur deposition effects. Eight case study areas and two 

supplemental study areas (Rocky Mountain National Park and Little Rock Lake, Wisconsin) are 

summarized in the REA based on ecosystem characteristics, indicators, and ecosystem service 

information. Case studies selected for aquatic acidification effects were the Adirondack 

Mountains and Shenandoah National Park. Kane Experimental Forest in Pennsylvania and 

Hubbard Brook Experimental Forest in New Hampshire were selected as case studies for 

terrestrial acidification. Aquatic nutrient enrichment case study locations were selected in the 

Potomac River Basin upstream of Chesapeake Bay and the Neuse River Basin upstream of the 
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Pamlico Sound in North Carolina. The CSS communities in southern California and the MCF 

communities in the San Bernardino and Sierra Nevada Mountains of California were selected as 

case studies for terrestrial nutrient enrichment. Two supplemental areas were also chosen, one in 

Rocky Mountain National Park for terrestrial nutrient enrichment and one in Little Rock Lake, 

Wisconsin for aquatic nutrient enrichment.

2. Key findings

In summary, based on case study analyses, the REA concludes that known or anticipated 

adverse ecological effects are occurring under current conditions and further concludes that these 

adverse effects continue into the future. Key findings from the air quality analyses, acidification 

and nutrient enrichment case studies, as well as general conclusions from evaluating additional 

welfare effects, are summarized below.

a. Air quality analyses

The air quality analyses in the REA encompass the current emissions sources of nitrogen 

and sulfur, as well as atmospheric concentrations, estimates of deposition of total nitrogen, 

policy-relevant background, and non-atmospheric loadings of nitrogen and sulfur to ecosystems, 

both nationwide and in the case study areas. Spatial fields of deposition were created using wet 

deposition measurements from the NADP National Trends Network and dry deposition 

predictions from the 2002 CMAQ model simulation. Some key conclusions from this analysis 

are:

(1) Total reactive nitrogen deposition and sulfur deposition are much greater in the East 

compared to most areas of the West.

(2) These regional differences in deposition correspond to the regional differences in oxides of 

nitrogen and SO2 concentrations and emissions, which are also higher in the East. Oxides of 
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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 subregional (e.g., the upper Midwest and Plains states). The relative amounts of oxidized versus 

reduced nitrogen deposition are consistent with the relative amounts of oxides of nitrogen

and NH3 emissions. Oxidized nitrogen deposition exceeds reduced nitrogen deposition in 

most of the case study areas; the major exception being the Neuse River/Neuse River Estuary 

Case Study Area.

(3) Reduced nitrogen deposition exceeds oxidized nitrogen deposition in the vicinity of local 

sources of NH3.

(4) There can be relatively large spatial variations in both total reactive nitrogen deposition and 

sulfur deposition within a case study area; this occurs particularly in those areas that contain 

or are near a high emissions source of oxides of nitrogen, NH3 and/or SO2.

(5) The seasonal patterns in deposition differ between the case study areas. For the case study

areas in the East, the season with the greatest amounts of total reactive nitrogen deposition 

correspond to the season with the greatest amounts of sulfur deposition. Deposition peaks in 

spring in the Adirondack, Hubbard Brook Experimental Forest, and Kane Experimental 

Forest case study areas, and it peaks in summer in the Potomac River/Potomac Estuary, 

Shenandoah, and Neuse River/Neuse River Estuary case study areas. For the case study 

areas in the West, there is less consistency in the seasons with greatest total reactive nitrogen 

and sulfur deposition in a given area. In general, both nitrogen and/or sulfur deposition 

peaks in spring or summer. The exception to this is the Sierra Nevada Range portion of the 

MCF Case Study Area, in which sulfur deposition is greatest in winter.

b. Deposition-related aquatic acidification
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The role of aquatic acidification in two eastern United States areas—northeastern New 

York’s Adirondack area and the Shenandoah area in Virginia—was analyzed in the REA to 

assess surface water trends in SO4

2-

and NO3

-

concentrations and ANC levels and to affirm the 

understanding that reductions in deposition could influence the risk of acidification. Monitoring 

data from the EPA-administered TIME)/ LTM programs and the EMAP were assessed for the 

years 1990 to 2006, and past, present and future water quality levels were estimated using both 

steady-state and dynamic biogeochemical models. 

Although wet deposition rates for SO2 and oxides of nitrogen in the Adirondack Case 

Study Area have reduced since the mid-1990s, current concentrations are still well above preacidification (1860) conditions. The MAGIC modeling predicts NO3

-

and SO4

2-

are 17- and 5-

fold higher today, respectively. The estimated average ANC for 44 lakes in the Adirondack Case 

Study Area is 62.1 μeq/L (± 15.7 μeq/L); 78 percent of all monitored lakes in the Adirondack 

Case Study Area have a current risk of Elevated, Severe, or Acute. Of the 78 percent, 31 percent

experience episodic acidification, and 18 percent are chronically acidic today.

(1) Based on the steady-state critical load model for the year 2002, 18 percent, 28 percent, 44

percent, and 58 percent of 169 modeled lakes received combined total sulfur and nitrogen 

deposition that exceeded critical loads corresponding to ANC limits of 0, 20, 50, and 100 

μeq/L respectively.

(2) Based on a deposition scenario that maintains current emission levels to 2020 and 2050, the 

simulation forecast indicates no improvement in water quality in the Adirondack Case Study 

Area. The percentage of lakes within the Elevated to Acute Concern classes remains the 

same in 2020 and 2050.
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(3) Since the mid-1990s, streams in the Shenandoah Case Study Area have shown slight declines 

in NO3 and SO4

2-

concentrations in surface waters. The ANC levels increased from about 50 

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. The 

MAGIC modeling predicts surface water concentrations of NO3 and SO4

2-

are 10- and 32-

fold higher today, respectively. The estimated average ANC for 60 streams in the 

Shenandoah Case Study Area is 57.9 μeq/L (± 4.5 μeq/L). Fifty-five percent of all monitored 

streams in the Shenandoah Case Study Area have a current risk of Elevated, Severe, or 

Acute. Of the 55 percent, 18 percent experience episodic acidification, and 18 percent are 

chronically acidic today.

(4) Based on the steady-state critical load model for the year 2002, 52 percent, 72 percent, 85

percent and 93 percent of 60 modeled streams received combined total sulfur and nitrogen 

deposition that exceeded critical loads corresponding to ANC limits of 0, 20, 50, and 100 

μeq/L respectively.

(5) Based on a deposition scenario that maintains current emission levels to 2020 and 2050, the 

simulation forecast indicates that a large number of streams would still have Elevated to 

Acute problems with acidity.

c. Deposition-related terrestrial acidification

The role of terrestrial acidification was examined in the REA using a critical load 

analysis for sugar maple and red spruce forests in the eastern U.S. by using the BC/Al ratio in 

acidified forest soils as an indicator to assess the impact of nitrogen and sulfur deposition on tree 

health. These are the two most commonly studied species in North America for impacts of 

acidification. At a BC/Al ratio of 1.2, red spruce growth can be reduced by 20 percent. Sugar 
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maple growth can be reduced by 20 percent at a BC/Al ratio of 0.6. Key findings of the case 

study are summarized below.

(1) Case study results suggest that the health of at least a portion of the sugar maple and red 

spruce growing in the U.S. may have been compromised with acidifying total nitrogen and 

sulfur deposition in 2002. The 2002 CMAQ/NADP total nitrogen and sulfur deposition 

levels exceeded three selected critical loads in 3 percent to 75 percent of all sugar maple plots 

across 24 states. The three critical loads ranged from 6,008 to 107 eq/ha/yr for the BC/Al 

ratios of 0.6, 1.2, and 10.0 (increasing levels of tree protection). The 2002 CMAQ/NADP 

total nitrogen and sulfur deposition levels exceeded three selected critical loads in 3 percent

to 36 percent of all red spruce plots across eight states. The three critical loads ranged from 

4,278 to 180 eq/ha/yr for the Bc/Al ratios of 0.6, 1.2, and 10.0 (increasing levels of tree 

protection).

(2) The SMB model assumptions made for base cation weathering (Bcw) and forest soil ANC 

input parameters are the main sources of uncertainty since these parameters are rarely 

measured and require researchers to use default values.

(3) The pattern of case study results suggests that nitrogen and sulfur acidifying deposition in the 

sugar maple and red spruce forest areas studied were similar in magnitude to the critical 

loads for those areas and both ecosystems are likely to be sensitive to any future changes in 

the levels of deposition.

d. Deposition-related aquatic nutrient enrichment

The role of nitrogen deposition in two main stem rivers feeding their respective estuaries 

was analyzed in the REA to determine if decreases in deposition could influence the risk of 

eutrophication as predicted using the ASSETS EI scoring system in tandem with SPARROW 
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modeling. This modeling approach provides a transferrable, intermediate-level analysis of the 

linkages between atmospheric deposition and receiving waters, while providing results on which 

conclusions could be drawn. A summary of findings follows:

(1) The 2002 CMAQ/NADP results showed that an estimated 40,770,000 kilograms (kg) of total 

nitrogen was deposited in the Potomac River watershed. The SPARROW modeling 

predicted that 7,380,000 kg N/yr of the deposited nitrogen reached the estuary (20 percent of 

the total load to the estuary). The overall ASSETS EI for the Potomac River and Potomac 

Estuary was Bad (based on all sources of N). 

(2) To improve the Potomac River and Potomac Estuary ASSETS EI score from Bad to Poor, a 

decrease of at least 78 percent in the 2002 total nitrogen atmospheric deposition load to the 

watershed would be required.

(3) The 2002 CMAQ/NADP results showed that an estimated 18,340,000 kg of total nitrogen 

was deposited in the Neuse River watershed. The SPARROW modeling predicted that 

1,150,000 kg N/yr of the deposited nitrogen reached the estuary (26 percent of the total load 

to the estuary). The overall ASSETS EI for the Neuse River/Neuse River Estuary was Bad.

(4) It was found that the Neuse River/Neuse River Estuary ASSETS EI score could not be 

improved from Bad to Poor with decreases only in the 2002 atmospheric deposition load to 

the watershed. Additional reductions would be required from other nitrogen sources within 

the watershed.

The small effect of decreasing atmospheric deposition in the Neuse River watershed is 

because the other nitrogen sources within the watershed are more influential than atmospheric 

deposition in affecting the total nitrogen loadings to the Neuse River Estuary, as estimated with 

the SPARROW model. A water body’s response to nutrient loading depends on the magnitude 
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(e.g., agricultural sources have a higher influence in the Neuse than in the Potomac), spatial 

distribution, and other characteristics of the sources within the watershed; therefore a reduction 

in nitrogen deposition does not always produce a linear response in reduced load to the estuary, 

as demonstrated by these two case studies.

e. Deposition-related terrestrial nutrient enrichment

California CSS and MCF communities were the focus of the Terrestrial Nutrient 

Enrichment Case Studies of the REA. Geographic information systems analysis supported a 

qualitative review of past field research to identify ecological benchmarks associated with CSS 

and mycorrhizal communities, as well as MCF’s nutrient-sensitive acidophyte lichen 

communities, fine-root biomass in Ponderosa pine and leached nitrate in receiving waters. These 

benchmarks, ranging from 3.1 to 17 kg N/ha/yr, were compared to 2002 CMAQ/NADP data to 

discern any associations between atmospheric deposition and changing communities. Evidence 

supports the finding that nitrogen alters CSS and MCF. Key findings include the following:

(1) The 2002 CMAQ/NADP nitrogen deposition data show that the 3.3 kg N/ha/yr benchmark 

has been exceeded in more than 93 percent of CSS areas (654,048 ha). This suggests that 

such deposition is a driving force in the degradation of CSS communities. One potentially

confounding factor is the role of fire. Although CSS decline has been observed in the 

absence of fire, the contributions of deposition and fire to the CSS decline require further 

research. The CSS is fragmented into many small parcels, and the 2002 CMAQ/NADP 12-

km grid data are not fine enough to fully validate the relationship between CSS distribution, 

nitrogen deposition, and fire.

(2) The 2002 CMAQ/NADP nitrogen deposition data exceeds the 3.1 kg N/ha/yr benchmark in 

more than 38% (1,099,133 ha) of MCF areas, and nitrate leaching has been observed in 
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surface waters. Ozone effects confound nitrogen effects on MCF acidophyte lichen, and the 

interrelationship between fire and nitrogen cycling requires additional research.

f. Additional effects

Ecological effects have also been documented across the U.S. where elevated nitrogen 

deposition has been observed, including the eastern slope of the Rocky Mountains where shifts 

in dominant algal species in alpine lakes have occurred where wet nitrogen deposition was only

about 1.5 kg N/ha/yr. High alpine terrestrial communities have a low capacity to sequester 

nitrogen deposition, and monitored deposition exceeding 3 to 4 kg N/ha/yr could lead to 

community-level changes in plant species, lichens and mycorrhizae.

Additional welfare effects are documented, but examined less extensively, in the REA. 

These effects include qualitative discussions related to visibility and materials damage, such as 

corrosion, erosion, and soiling of paint and buildings which are being addressed in the PM

NAAQS review currently underway. A discussion of the causal relationship between sulfur 

deposition (as sulfate, SO4

2-

) and increased mercury methylation in wetlands and aquatic 

environments is also included in the REA. On this subject the REA concludes that decreases in 

SO4

2-

deposition will likely result in decreases in MeHg concentration; however, spatial and 

biogeochemical variations nationally hinder establishing large scale dose-response relationships. 

Several additional issues concerning oxides of nitrogen were addressed in the REA. 

Consideration was also given to N2O, a potent GHG. The REA concluded that it is most 

appropriate to analyze the role of N2O in the context of all of the GHGs rather than as part of the 

REA for this review. The REA considered nitrogen deposition and its correlation with the rate 

of photosynthesis and net primary productivity. Nitrogen addition ranging from 15.4 to 300 kg 

N/ha/yr is documented as increasing wetland N2O production by an average of 207 percent
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across all ecosystems. Nitrogen addition ranging from 30 to 240 kg N/ha/yr increased CH4

emissions by 115 percent, averaged across all ecosystems, and methane uptake was reduced by 

38 percent averaged across all ecosystems when nitrogen addition ranged from 10 to 560 kg 

N/ha/yr, but reductions were only significant for coniferous and deciduous forests. The 

heterogeneity of ecosystems across the U.S., however, introduces variations into dose-response 

relationships.

The phytotoxic effects of oxides of nitrogen and sulfur on vegetation were also briefly 

discussed in the REA which concluded that since a unique secondary NAAQS exists for SO2, 

and concentrations of nitric oxide (NO), NO2 and PAN are rarely high enough to have phytotoxic 

effects on vegetation, further assessment was not warranted at this time. 

3. Conclusions on Effects 

For aquatic and terrestrial acidification effects, a similar conceptual approach was used 

(critical loads) to evaluate the impacts of multiple pollutants on an ecological endpoint, whereas 

the approaches used for aquatic and terrestrial nutrient enrichment were fundamentally distinct. 

Although the ecological indicators for aquatic and terrestrial acidification (i.e., ANC and BC/Al) 

are very different, both ecological indicators are well-correlated with effects such as reduced 

biodiversity and growth. While aquatic acidification is clearly the targeted effect area with the 

highest level of confidence, the relationship between atmospheric deposition and an ecological 

indicator is also quite strong for terrestrial acidification. The main drawback with the 

understanding of terrestrial acidification is that the data are based on laboratory responses rather 

than field measurements. Other stressors that are present in the field but that are not present in 

the laboratory may confound this relationship.
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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 nonatmospheric sources of nitrogen and the influence of both oxidized and reduced forms of 

nitrogen, particularly in large watersheds and coastal areas. However, a strong relationship 

exists between atmospheric deposition of nitrogen and ecological effects in high alpine lakes in 

the Rocky Mountains because atmospheric deposition is the only source of nitrogen to these 

systems. There is also a strong weight-of-evidence regarding the relationships between 

ecological effects attributable to terrestrial nitrogen nutrient enrichment; however, ozone and 

climate change may be confounding factors. In addition, the response for other species or 

species in other regions of the U.S. has not been quantified.

C. Adversity of Effects to Public Welfare

Characterizing a known or anticipated adverse effect to public welfare is an important 

component of developing any secondary NAAQS. According to the CAA, welfare effects 

include: “effects on soils, water, crops, vegetation, manmade materials, animals, wildlife, 

weather, visibility, and climate, damage to and deterioration of property, and hazards to 

transportation, as well as effect on economic values and on personal comfort and well-being, 

whether caused by transformation, conversion, or combination with other air pollutants” (CAA, 
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Section 302(h)). While the text above lists a number of welfare effects, these effects do not 

define public welfare in and of themselves. 

Although there is no specific definition of adversity to public welfare, the paradigm of 

linking adversity to public welfare to disruptions in ecosystem structure and function has been 

used broadly by EPA to categorize effects of pollutants from the cellular to the ecosystem level. 

An evaluation of adversity to public welfare might consider the likelihood, type, magnitude, and 

spatial scale of the effect as well as the potential for recovery and any uncertainties relating to 

these considerations. 

Similar concepts were used in past reviews of secondary NAAQS for ozone and PM 

(relating to visibility), as well as in initial reviews of effects from lead deposition. Because 

oxides of nitrogen and sulfur are deposited from ambient sources into ecosystems where they

affect changes to organisms, populations and ecosystems, the concept of adversity to public 

welfare as a result of alterations in structure and function of ecosystems is an appropriate 

consideration for this review. 

Based on information provided in the PA, the following section discusses how ecological 

effects from deposition of oxides of nitrogen and sulfur relate to adversity to public welfare. In 

the PA, public welfare was discussed in terms of loss of ecosystem services (defined below), 

which in some cases can be monetized. Each of the four main effect areas (aquatic and terrestrial 

acidification and aquatic and terrestrial nutrient over-enrichment) are discussed including current 

ecological effects and associated ecosystem services.

1. Ecosystem Services 

The PA defines ecosystem services as the benefits individuals and organizations obtain 

from ecosystems. Ecosystem services can be classified as provisioning (food and water), 
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regulating (control of climate and disease), cultural (recreational, existence, spiritual, 

educational), and supporting (nutrient cycling). Conceptually, changes in ecosystem services 

may be used to aid in characterizing a known or anticipated adverse effect to public welfare. In 

the REA and PA ecosystem services are discussed as a method of assessing the magnitude and 

significance to the public of resources affected by ambient concentrations of oxides of nitrogen 

and sulfur and deposition in sensitive ecosystems. 

The EPA has in previous NAAQS reviews defined ecological goods and services for the 

purposes of a Regulatory Impact Analysis as the “outputs of ecological functions or processes 

that directly or indirectly contribute to social welfare or have the potential to do so in the future. 

Some outputs may be bought and sold, but most are not marketed.” It is especially important to

acknowledge that it is difficult to measure and/or monetize the goods and services supplied by 

ecosystems. It can be informative in characterizing adversity to public welfare to attempt to 

place an economic valuation on the set of goods and services that have been identified with 

respect to a change in policy however it must be noted that this valuation will be incomplete and 

illustrative only. 

Knowledge about the relationships linking ambient concentrations and ecosystem 

services is considered in the PA as one method by which to inform a policy judgment on a 

known or anticipated adverse public welfare effect. For example, a change in an ecosystem 

structure and process, such as foliar injury, would be classified as an ecological effect, with the 

associated changes in ecosystem services, such as primary productivity, food availability, forest 

products, and aesthetics (e.g., scenic viewing), classified as public welfare effects. Additionally, 

changes in biodiversity would be classified as an ecological effect, and the associated changes in 
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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 ANC) and then to changes in ecosystems. Appendix 8 of the 

REA links the changes in ecosystems to the services they provide (e.g., fish species richness and 

its influence on recreational fishing). To the extent possible for each targeted effect area, the 

REA linked ambient concentrations of nitrogen and sulfur (i.e., ambient air quality indicators) to 

deposition in sensitive ecosystems (i.e., exposure pathways), and then to system response as 

measured by a given ecological indicator (e.g., lake and stream acidification as measured by 

ANC). The ecological effect (e.g., changes in fish species richness) was then, where possible, 

associated with changes in ecosystem services and the corresponding public welfare effects (e.g., 

recreational fishing). 

2. Effects on Ecosystem Services

The process used to link ecological indicators to ecosystem services is discussed 

extensively in appendix 8 of the REA. In brief, for each case study area assessed, the ecological 

indicators are linked to an ecological response that is subsequently linked to associated services 

to the extent possible. For example, in the case study for aquatic acidification the chosen 

ecological indicator is ANC which can be linked to the ecosystem service of recreational fishing. 

Although recreational fishing losses are the only service effects that can be independently 

quantified or monetized at this time, there are numerous other ecosystem services that may be 

related to the ecological effects of acidification.
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While aquatic acidification is the focus of this proposed standard, the other effect areas 

were also analyzed in the REA and these ecosystems are being harmed by nitrogen and sulfur 

deposition and will obtain some measure of protection with any decrease in that deposition 

regardless of the reason for the decrease. The following summarizes the current levels of 

specific ecosystem services for aquatic and terrestrial acidification and aquatic and terrestrial 

nutrient over-enrichment and attempts to quantify and when possible monetize the harm to 

public welfare, as represented by ecosystem services, due to nitrogen and sulfur deposition. 

a. Aquatic Acidification

Acidification of aquatic ecosystems primarily affects the ecosystem services that are 

derived from the fish and other aquatic life found in surface waters. In the northeastern United 

States, the surface waters affected by acidification are not a major source of commercially raised 

or caught fish; however, they are a source of food for some recreational and subsistence fishers 

and for other consumers. Although data and models are available for examining the effects on 

recreational fishing, relatively little data are available for measuring the effects on subsistence 

and other consumers. Inland waters also provide aesthetic and educational services along with 

non-use services, such as existence value (protection and preservation with no expectation of 

direct use). In general, inland surface waters such as lakes, rivers, and streams also provide a 

number of regulating services, playing a role in hydrological regimes and climate regulation. 

There is little evidence that acidification of freshwaters in the northeastern U.S. has significantly

degraded these specific services; however, freshwater ecosystems also provide biological control 

services by providing environments that sustain delicate aquatic food chains. The toxic effects 

of acidification on fish and other aquatic life impair these services by disrupting the trophic 

structure of surface waters. Although it is difficult to quantify these services and how they are 
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affected by acidification, it is worth noting that some of these services may be captured through 

measures of provisioning and cultural services. For example, these biological control services 

may serve as “intermediate” inputs that support the production of “final” recreational fishing and 

other cultural services.

As summarized in Chapter 4 of the PA, recent studies indicate that acidification of lakes 

and streams can result in significant loss in economic value. For example, data indicate that 

more than 9 percent of adults in the northeastern part of the country participate annually in 

freshwater fishing yielding 140 million freshwater fishing days. Each fishing day has an 

estimated average value per day of $35. Therefore, the implied total annual value of freshwater 

fishing in the northeastern U.S. was $5 billion in 2006. Embedded in these numbers is a degree 

of harm to recreational fishing services due to acidification that has occurred over time. These 

harms have not been quantified on a regional scale; however, a case study was conducted in the 

Adirondacks area (US EPA, 2011, section 4.4.2). 

In the Adirondacks case study, estimates of changes in recreational fishing services were 

determined, as well as changes more broadly in “cultural” ecosystem services (including 

recreational, aesthetic, and nonuse services). First, the MAGIC model (US EPA, 2009,

Appendix 8 and section 2.2) was applied to 44 lakes to predict what ANC levels would be under 

both “business as usual” conditions (i.e., allowing for some decline in deposition due to existing 

regulations) and pre-emission (i.e., background) conditions. Second, to estimate the recreational 

fishing impacts of aquatic acidification in these lakes, an existing model of recreational fishing 

demand and site choice was applied. This model predicts how recreational fishing patterns in the 

Adirondacks would differ and how much higher the average annual value of recreational fishing 

services would be for New York residents if lake ANC levels corresponded to background 
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(rather than business as usual) conditions. To estimate impacts on a broader category of cultural 

(and some provisioning) ecosystem services, results from the Banzhaf et al (2006) valuation 

survey of New York residents were adapted and applied to this context. The survey used a 

contingent valuation approach to estimate the average annual household willingness to pay

(WTP) for future reductions in the percent of Adirondack lakes impaired by acidification. The 

focus of the survey was on impacts on aquatic resources. Pretesting of the survey indicated that 

respondents nonetheless tended to assume that benefits would occur in the condition of birds and 

forests as well as in recreational fishing. 

By extrapolating the 44 lake Adirondack case study to all 3,000 Adirondack lakes and by 

applying the WTP survey results to all New York residents, the study estimated aggregated 

benefits between $300 and $800 million annually for the equivalent of improving lakes in the 

Adirondacks region to an ANC level of 50 μeq/L. The REA estimated 44 percent 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.

b. Terrestrial Acidification
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Chapters 4.4.3 and 4.4.4 of the PA review several economic studies of areas sensitive to 

terrestrial acidification. Forests in the northeastern U.S. provide several important and valuable 

provisioning ecosystem services, which are reflected in the production and sales of tree products.

Sugar maples are a particularly important commercial hardwood tree species in the United 

States, producing timber and maple syrup that provide hundreds of millions of dollars in 

economic value annually. Red spruce is also used in a variety of wood products and provides up 

to $100 million in economic value annually. Although the data do not exist to directly link 

acidification damages to economic values of lost recreational ecosystem services in forests, these 

resources are valuable to the public. A recent study, reviewed in the PA, suggests that the total 

annual value of recreational off-road driving was more than $9 billion and the value of hunting 

and wildlife viewing was more than $4 billion each in the northeastern States. The EPA is not 

able to quantify at this time the specific effects on these values of acid deposition, or of any

specific reductions in deposition, relative to the effects of many other factors that may affect 

them.

c. Nutrient Enrichment

Chapters 4.4.5 and 4.4.6 of the PA summarize economic studies of east coast estuaries 

affected by nutrient over-enrichment or eutrophication. Estuaries in the eastern United States are 

important for fish and shellfish production. The estuaries are capable of supporting large stocks 

of resident commercial species, and they serve as the breeding grounds and interim habitat for 

several migratory species. To provide an indication of the magnitude of provisioning services 

associated with coastal fisheries, from 2005 to 2007, the average value of total catch was $1.5 

billion per year in 15 East Coast states. Estuaries also provide an important and substantial 

variety of cultural ecosystem services, including water-based recreational and aesthetic services. 
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For example, data indicate that 4.8 percent of the population in coastal states from North 

Carolina to Massachusetts participated in saltwater fishing, with a total of 26 million saltwater 

fishing days in 2006. Based on estimates in the PA, total recreational value from these saltwater 

fishing days was approximately $1.3 billion. Recreational participation estimates for 1999–2000 

showed almost 6 million individuals participated in motorboating in coastal states from North 

Carolina to Massachusetts. The aggregate value of these coastal motorboating outings was $2 

billion per year. EPA is not able to quantify at this time the specific effects on these values of 

nitrogen deposition, or of any specific reductions in deposition, relative to the effects of many 

other factors that may affect them.

Terrestrial ecosystems can also suffer from nutrient over-enrichment. Each ecosystem is 

different in its composition of species and nutrient requirements. Changes to individual 

ecosystems from changes in nitrogen deposition can be hard to assess economically. Relative 

recreational values are often determined by public use information. Chapter 4.4.7 of the PA

reviewed studies related to park use in California. Data from California State Parks indicate that 

in 2002, 68.7 percent of adult residents participated in trail hiking for an average of 24.1 days per 

year. The analyses in the PA indicate that the aggregate annual benefit for California residents 

from trail hiking in 2007 was $11.59 billion. EPA is not able to quantify at this time the specific 

effects on these values of nitrogen deposition, or of any specific reductions in deposition, relative 

to the effects of many other factors that may affect them.

The PA also identified fire regulation as a service that could be affected by nutrient overenrichment of the CSS and MCF ecosystems by encouraging growth of more flammable grasses, 

increasing fuel loads, and altering the fire cycle. Over the 5-year period from 2004 to 2008, 

Southern California experienced, on average, over 4,000 fires per year, burning, on average, over 
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400,000 acres per year. It is not possible at this time to quantify the contribution of nitrogen 

deposition, among many other factors, to increased fire risk.

3. Summary

Adversity to public welfare can be understood by looking at how deposition of oxides of 

nitrogen and sulfur affect the ecological functions of an ecosystem (see II.A.), and then

understanding the ecosystem services that are degraded. The monetized value of the ecosystem 

services provided by ecosystems that are sensitive to deposition of oxides of nitrogen and sulfur 

are in the billions of dollars each year, though it is not possible to quantify or monetize at this 

time the effects on these values of nitrogen and sulfur deposition or of any changes in deposition 

that may result from new secondary standards. Many lakes and streams are known to be 

degraded by acidic deposition which affects recreational fishing and tourism. Forest growth is 

likely suffering from acidic deposition in sensitive areas affecting red spruce and sugar maple 

timber production, sugar maple syrup production, hiking, aesthetic enjoyment and tourism. 

Nitrogen deposition contributes significantly to eutrophication in many estuaries affecting fish 

production, swimming, boating, aesthetic enjoyment and tourism. Ecosystem services are likely 

affected by nutrient enrichment in many natural and scenic terrestrial areas, affecting 

biodiversity, including habitat for rare and endangered species, fire control, hiking, aesthetic 

enjoyment and tourism. 

D. Adequacy of the Current Standards

An important issue to be addressed in the current review of the secondary standards for 

oxides of nitrogen and sulfur is whether, in view of the scientific evidence reflected in the ISA, 

additional information on exposure and risk discussed in the REA, and conclusions drawn from 

the PA, the existing standards provide adequate protection. The Administrator therefore, has 
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considered the extent to which the current standards are adequate for the protection of public 

welfare. Having reached the general conclusion that aquatic and terrestrial ecosystems can be 

degraded by deposition of oxides of nitrogen and sulfur, it is then necessary to first evaluate the 

appropriateness (in terms of form and structure) of the current standards to address the ecological 

effects of oxides of nitrogen and sulfur as well as the adequacy of the current secondary

standards for oxides of nitrogen and sulfur to provide requisite protection by considering to what 

degree risks to sensitive ecosystems would be expected to occur in areas that meet the current 

standards. Conclusions regarding the adequacy of the current standards are based on the 

available ecological effects, exposure and risk-based evidence. In evaluating the strength of this 

information, EPA has taken into account the uncertainties and limitations in the scientific 

evidence. This section addresses the adequacy of the current standards to protect against direct 

exposure effects on plants from oxides of nitrogen and sulfur, the appropriateness of the current 

structure of the standards to address deposition-related effects of oxides of nitrogen and sulfur on 

sensitive ecosystems and finally, the adequacy of such standards to protect against adverse 

effects related to the deposition of oxides of nitrogen and sulfur. 

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 standard3, i.e., an annual standard set for NO2 to protect against 

adverse effects on vegetation from direct exposure to ambient oxides of nitrogen. For oxides of 

sulfur, the current secondary standard is a 3-hour standard intended to provide protection for 

plants from the direct foliar damage associated with atmospheric concentrations of SO2. It is 

 

3 The current primary NO2 standard has recently been changed to the 3-year average of the 98th percentile of the 

annual distribution of the 1 hour daily maximum of the concentration of NO2. The current secondary standard 

remains as it was set in 1971.
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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. 

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

This section addresses two concepts necessary to evaluate the current standards in the 

context of deposition related effects. First, appropriateness of the current standards is considered 

with regard to indicator, form, level and averaging time. This discussion centers around the 

ability of the current standards to evaluate and provide protection against deposition related 

effects that vary spatially and temporally. It includes particular emphasis on the indicators and 

forms of the current standards and the degree to which they are ecologically relevant with regard 

to deposition related effects. Second, this section evaluates the current standards in terms of

adequacy of protection. 

a. Appropriateness

The ISA has established that the major effects of concern for this review of the oxides of 

nitrogen and sulfur standards are associated with deposition of nitrogen and sulfur caused by 

atmospheric concentrations of oxides of nitrogen and sulfur. The current standards are not 

directed toward depositional effects, and none of the elements of the current NAAQS – indicator, 
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form, averaging time, and level – are suited for addressing the effects of nitrogen and sulfur

deposition. 

Five issues arise that call into question the ecological relevance of the structure of the 

current secondary standards for oxides of nitrogen and sulfur. 

(1) The current SO2 secondary standard (0.5 ppm SO2 over a 3-hour average) does not utilize an 

exposure period that is relevant for ecosystem impacts. The majority of deposition related 

impacts are associated with depositional loads that occur over periods of months to years. 

This differs significantly from exposures associated with hourly concentrations of SO2 as 

measured by the current secondary standard. By addressing short-term concentrations, the 

current SO2 secondary standard, while protective against direct foliar effects from gaseous 

oxides of sulfur, does not take into account the findings of effects in the ISA, which notes the 

relationship between annual deposition of sulfur and acidification effects which are likely to 

be more severe and widespread than phytotoxic effects under current ambient conditions, and 

include effects from long term deposition as well as short term. Acidification is a process 

that occurs over time because the ability of an aquatic system to counteract acidic inputs is 

reduced as natural buffers are used more rapidly than they can be replaced through geologic 

weathering. The relevant period of exposure for ecosystems is, therefore, not the exposures 

captured in the short averaging time of the current SO2 secondary standard. The current 

secondary standard for oxides of nitrogen is an annual standard (0.053 ppm averaged over 1 

year) and as such is more ecologically relevant.

(2) Current standards do not utilize appropriate atmospheric indicators. Nitrogen dioxide and 

SO2 are used as the component of oxides of nitrogen and sulfur that are measured, but they 

do not provide a complete link to the direct effects on ecosystems from deposition of oxides 
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of nitrogen and sulfur as they do not capture all relevant chemical species of oxidized 

nitrogen and oxidized sulfur that contribute to deposition. The ISA provides evidence that 

deposition related effects are linked with total nitrogen and total sulfur deposition, and thus 

all forms of oxidized nitrogen and oxidized sulfur that are deposited will contribute to effects 

on ecosystems. Thus, by using atmospheric NO2 and SO2 concentrations as indicators, the 

current standards address only a fraction of total atmospheric oxides of nitrogen and sulfur, 

and do not take into account the effects from deposition of total atmospheric oxides of 

nitrogen and sulfur. This suggests that more comprehensive atmospheric indicators should 

be considered in designing ecologically relevant standards. 

(3) Current standards reflect separate assessments of the two individual pollutants, NO2 and SO2, 

rather than assessing the joint impacts of deposition to ecosystems. Recognizing the role that 

each pollutant plays in jointly affecting ecosystem indicators, functions, and services is vital 

to developing a meaningful standard. The clearest example of this interaction is in 

assessment of the impacts of acidifying deposition on aquatic ecosystems. Acidification in 

an aquatic ecosystem depends on the total acidifying potential of the deposition of both 

nitrogen and sulfur from both atmospheric deposition of oxides of nitrogen and sulfur as well 

as the inputs from other sources of nitrogen and sulfur such as reduced nitrogen and nonatmospheric 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 
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nitrogen and sulfur and reduced forms of nitrogen, which also contribute to acidification and 

nutrient enrichment. It is important that the structure of the standards address the role of 

reduced nitrogen in determining the ecological effects resulting from deposition of 

atmospheric oxides of nitrogen and sulfur. Consideration will also have to be given to total 

loadings as ecosystems respond to all sources of nitrogen and sulfur.

(4) Current standards do not take into account variability in ecosystem sensitivity. Ecosystems 

are not uniformly distributed either spatially or temporally in their sensitivity to oxides of 

nitrogen and sulfur. Therefore, failure to account for the major determinants of variability, 

including geological and soil characteristics related to the sensitivity to acidification or 

nutrient enrichment as well as atmospheric and landscape characteristics that govern rates of 

deposition, may lead to standards that do not provide requisite levels of protection across 

ecosystems. The current structures of the standards do not address the complexities in the 

responses of ecosystems to deposition of oxides of nitrogen and sulfur. Ecosystems contain 

complex groupings of organisms that respond in various ways to the alterations of soil and 

water that result from deposition of nitrogen and sulfur compounds. Different ecosystems

therefore respond in different ways depending on a multitude of factors that control how 

deposition is integrated into the system. For example, the same levels of deposition falling 

on limestone dominated soils have a very different effect from those falling on shallow 

glaciated soils underlain with granite. One system may over time display no obvious 

detriment while the other may experience a catastrophic loss in fish communities. This 

degree of sensitivity is a function of many atmospheric factors that control rates of deposition 

as well as ecological factors that control how an ecosystem responds to that deposition. The 

current standards do not take into account spatial and seasonal variations, not only in 
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depositional loadings, but also in sensitivity of ecosystems exposed to those loadings. Based 

on the discussion summarized above, the PA concludes that the current secondary standards 

for oxides of nitrogen and oxides of sulfur are not ecologically relevant in terms of averaging 

time, form, level or indicator.

b. Adequacy of Protection

As described in the PA, ambient conditions in 2005 indicate that the current SO2 and NO2

secondary standards were not exceeded at that time (US EPA, 2011, Figures 6-1 and 6-2) in 

locations where negative ecological effects have been observed. In many locations, SO2 and 

NO2 concentrations are substantially below the levels of the secondary standards. This pattern 

suggests that levels of deposition and any negative effects on ecosystems due to deposition of 

oxides of nitrogen and sulfur under recent conditions are occurring even though areas meet or are 

below current standards. In addition, based on conclusions in the REA, these levels will not 

decline in the future to levels below which it is reasonable to anticipate effects.

In determining the adequacy of the current secondary standards for oxides of nitrogen 

and sulfur the PA considered the extent to which ambient deposition contributes to loadings in 

ecosystems. Since the last review of the secondary standard for oxides of nitrogen, a great deal 

of information on the contribution of atmospheric deposition associated with ambient oxides of 

nitrogen has become available. The REA presents a thorough assessment of the contribution of 

oxidized nitrogen to nitrogen deposition throughout the U.S., and the relative contributions of 

ambient oxidized and reduced forms of nitrogen. The REA concludes that based on that 

analysis, ambient oxides of nitrogen are a significant component of atmospheric nitrogen 

deposition, even in areas with relatively high rates of deposition of reduced nitrogen. In 
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addition, atmospheric deposition of oxidized nitrogen contributes significantly to total nitrogen 

loadings in nitrogen sensitive ecosystems. 

The ISA summarizes the available studies of relative nitrogen contribution and finds that 

in much of the U.S., oxides of nitrogen contribute from 50 to 75 percent of total atmospheric 

deposition relative to total reactive nitrogen, which includes oxidized and reduced nitrogen 

species (US EPA, 2008, section 2.8.4). Although the proportion of total nitrogen loadings 

associated with atmospheric deposition of nitrogen varies across locations, the ISA indicates that 

atmospheric nitrogen deposition is the main source of new anthropogenic nitrogen to most 

headwater streams, high elevation lakes, and low-order streams. Atmospheric nitrogen

deposition contributes to the total nitrogen load in terrestrial, wetland, freshwater and estuarine 

ecosystems that receive nitrogen through multiple pathways. In several large estuarine systems, 

including the Chesapeake Bay, atmospheric deposition accounts for between 10 and 40 percent 

of total nitrogen loadings (US EPA, 2008). 

Atmospheric concentrations of oxides of sulfur account for nearly all sulfur deposition in 

the US. For the period 2004-2006, mean sulfur deposition in the U.S. was greatest east of the 

Mississippi River with the highest deposition amount, 21.3 kg S/ha-yr, in the Ohio River Valley 

where most recording stations reported 3-year averages >10 kg S/ha-yr. Numerous other stations 

in the East reported S deposition >5 kg S/ha-yr. Total sulfur deposition in the U.S. west of the 

100th meridian was relatively low, with all recording stations reporting <2 kg S/ha-yr and many 

reporting <1 kg S/ha-yr. Sulfur was primarily deposited in the form of wet SO4

2−

followed in 

decreasing order by a smaller proportion of dry SO2 and a much smaller proportion of deposition 

as dry SO4

2−

. 
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As discussed throughout the REA (US EPA, 2009 and section II.B above), there are 

several key areas of risk that are associated with ambient concentrations of oxides of nitrogen 

and sulfur. As noted earlier, in previous reviews of the secondary standards for oxides of 

nitrogen and sulfur, the standards were designed to protect against direct exposure of plants to 

ambient concentrations of the pollutants. A significant shift in understanding of the effects of 

oxides of nitrogen and sulfur has occurred since the last reviews, reflecting the large amount of 

research that has been conducted on the effects of deposition of nitrogen and sulfur to 

ecosystems. The most significant current risks of adverse effects to public welfare are those 

related to deposition of oxides of nitrogen and sulfur to both terrestrial and aquatic ecosystems.

These risks fall into two categories, acidification and nutrient enrichment, which were 

emphasized in the REA as most relevant to evaluating the adequacy of the existing standards in 

protecting public welfare from adverse ecological effects.

i. Aquatic acidification

The focus of the REA case studies was on determining whether deposition of sulfur and 

oxidized nitrogen in locations where ambient oxides of nitrogen and sulfur were at or below the 

current standards was resulting in acidification and related effects, including episodic 

acidification and mercury methylation. Based on the case studies conducted for lakes in the 

Adirondacks and streams in Shenandoah National Park (case studies are discussed more fully in 

section II.B and US EPA, 2009), there is significant risk to acid sensitive aquatic ecosystems at 

atmospheric concentrations of oxides of nitrogen and sulfur at or below the current standards. 

The REA also supports strongly a relationship between atmospheric deposition of oxides of 

nitrogen and sulfur and loss of ANC in sensitive ecosystems and indicates that ANC is an 

excellent indicator of aquatic acidification. The REA also concludes that at levels of deposition 
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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. 

In describing the effects of acidification in the two case study areas the REA uses the 

approach of describing benchmarks in terms of ANC values. Many locations in sensitive areas 

of the U.S. have ANC levels below benchmark levels for ANC classified as severe, elevated, or 

moderate concern (US EPA, 2011, Figure 2-1). The average current ANC levels across 44 lakes 

in the Adirondack case study area is 62.1 μeq/L (moderate concern). However, 44 percent of 

lakes had deposition levels exceeding the critical load for an ANC of 50 μeq/L (elevated), and 28 

percent of lakes had deposition levels exceeding the (higher) critical load for an ANC of 20 

μeq/L (severe) (US EPA, 2009, section 4.2.4.2). This information indicates that almost half of 

the 44 lakes in the Adirondacks case study area are at an elevated concern level, and almost a 

third are at a severe concern level. These levels are associated with greatly diminished fish 

species diversity, and losses in the health and reproductive capacity of remaining populations. 

Based on assessments of the relationship between number of fish species and ANC level in both 
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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.

Many streams in the Shenandoah case study area also have levels of deposition that are 

associated with ANC levels classified as severe, elevated, or moderate concern. The average

ANC under recent conditions for the 60 streams evaluated in the Shenandoah case study area is 

57.9 μeq/L, indicating moderate concern. However, 85 percent of these streams had recent 

deposition exceeding the critical load for an ANC of 50 μeq/L, and 72 percent exceeded the 

critical load for an ANC of 20 μeq/L. As with the Adirondacks area, this information suggests 

that ANC levels may decline in the future and significant numbers of sensitive streams in the 

Shenandoah area are at risk of adverse impacts on fish populations if recent conditions persist. 

Many other streams in the Shenandoah area are also likely to experience conditions of elevated 

to severe concern based on the prevalence in the area of bedrock geology associated with 

increased sensitivity to acidification suggesting that effects due to stream acidification could be 

widespread in the Shenandoah area (US EPA, 2009, section 4.2.6.2).

In addition to these chronic acidification effects, the ISA notes that “consideration of 

episodic acidification greatly increases the extent and degree of estimated effects for acidifying 

deposition on surface waters” (US EPA, 2008, section 3.2.1.6). Some studies show that the 

number of lakes that could be classified as acid-impacted based on episodic acidification is 2 to 3 

times the number of lakes classified as acid-impacted based on chronic ANC. These episodic 
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acidification events can have long term effects on fish populations (US EPA, 2008, section 

3.2.1.6). Under recent conditions, episodic acidification has been observed in locations in the 

eastern U.S. and in the mountainous western U.S. (US EPA, 2008, section 3.2.1.6). 

The ISA, REA and PA all conclude that the current standards are not adequate to protect 

against the adverse impacts of aquatic acidification on sensitive ecosystems. A recent survey, as 

reported in the ISA, found sensitive streams in many locations in the U.S., including the 

Appalachian Mountains, the Coastal Plain, and the Mountainous West (US EPA, 2008, section 

4.2.2.3). In these sensitive areas, between 1 and 6 percent of stream kilometers are chronically 

acidified. The REA further concludes that both the Adirondack and Shenandoah case study

areas are currently receiving deposition from ambient oxides of nitrogen and sulfur in excess of 

their ability to neutralize such inputs. In addition, based on the current emission scenarios, 

forecast modeling out to the year 2020 as well as 2050 indicates a large number of streams in 

these areas will still be adversely impacted (section II.B). Based on these considerations, the PA

concludes that the current secondary NAAQS for oxides of nitrogen and sulfur do not provide 

adequate protection of sensitive ecosystems with regard to aquatic acidification. 

ii. Terrestrial acidification

Based on the terrestrial acidification case studies, Kane Experimental Forest in 

Pennsylvania and Hubbard Brook Experimental Forest described in section II.B) of sugar maple 

and red spruce habitat, the REA concludes that there is significant risk to sensitive terrestrial 

ecosystems from acidification at atmospheric concentrations of NO2 and SO2 at or below the 

current standards. The ecological indicator selected for terrestrial acidification is the BC/Al, 

which has been linked to tree health and growth. The results of the REA strongly support a 

relationship between atmospheric deposition of oxides of nitrogen and sulfur and BC/Al, and that 
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BC/Al is a good indicator of terrestrial acidification. At levels of deposition associated with 

oxides of nitrogen and sulfur concentrations at or below the current standards, BC/Al levels are 

expected to be below benchmark values that are associated with significant effects on tree health 

and growth. Such degradation of terrestrial ecosystems could affect ecosystem services such as 

habitat provisioning, endangered species, goods production (timber, syrup, etc.) among others. 

Many locations in sensitive areas of the U.S. have BC/Al levels below benchmark levels 

classified as providing low to intermediate levels of protection to tree health. At a BC/Al ratio of 

1.2 (intermediate level of protection), red spruce growth can be reduced by 20 percent. At a 

BC/Al ratio of 0.6 (low level of protection), sugar maple growth can be decreased by 20 percent. 

The REA did not evaluate broad sensitive regions. However, in the sugar maple case study area 

(Kane Experimental Forest), recent deposition levels are associated with a BC/Al ratio below 

1.2, indicating between intermediate and low level of protection, which would indicate the 

potential for a greater than 20 percent reduction in growth. In the red spruce case study area 

(Hubbard Brook Experimental Forest), recent deposition levels are associated with a BC/Al ratio 

slightly above 1.2, indicating slightly better than an intermediate level of protection (US EPA, 

2009, section 4.3.5.1). 

Over the full range of sugar maple, 12 percent of evaluated forest plots exceeded the 

critical loads for a BC/Al ratio of 1.2, and 3 percent exceeded the critical load for a BC/Al ratio 

of 0.6. However, there was large variability across states. In New Jersey, 67 percent of plots 

exceeded the critical load for a BC/Al ratio of 1.2, while in several states on the outskirts of the 

range for sugar maple (e.g. Arkansas, Illinois) no plots exceeded the critical load for a BC/Al

ratio of 1.2. For red spruce, overall 5 percent of plots exceeded the critical load for a BC/Al ratio 

of 1.2, and 3 percent exceeded the critical load for a BC/Al ratio of 0.6. In the major red spruce 
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producing states (Maine, New Hampshire, and Vermont), critical loads for a BC/Al ratio of 1.2 

were exceeded in 0.5, 38, and 6 percent of plots, respectively.

The ISA, REA and PA all conclude that the current standards are not adequate to protect 

against the adverse impacts of terrestrial acidification on sensitive ecosystems. As stated in the 

REA and PA, the main drawback, with the understanding of terrestrial acidification lies in the 

sparseness of available data by which we can predict critical loads and that the data are based on 

laboratory responses rather than field measurements. Other stressors that are present in the field 

but that are not present in the laboratory may confound this relationship. The REA does 

however, conclude that the case study results, when extended to a 27 state region, show that 

nitrogen and sulfur acidifying deposition in the sugar maple and red spruce forest areas caused

the calculated Bc/Al ratio to fall below 1.2 (the intermediate level of protection) in 12 percemt of 

the sugar maple plots and 5 percent of the red spruce plots; however, results from individual 

states ranged from 0 to 67 percent of the plots for sugar maple and 0 to 100 percent of the plots 

for red spruce.

iii. Terrestrial nutrient enrichment

Nutrient enrichment effects are due to nitrogen loadings from both atmospheric and nonatmospheric sources. Evaluation of nutrient enrichment effects requires an understanding that 

nutrient inputs are essential to ecosystem health and that specific long term levels of nutrients in 

a system affect the types of species that occur over long periods of time. Short term additions of 

nutrients can affect species competition, and even small additions of nitrogen in areas that are 

traditionally nutrient poor can have significant impacts on productivity as well as species 

composition. Most ecosystems in the U.S. are nitrogen-limited, so regional decreases in 

emissions and deposition of airborne nitrogen compounds could lead to some decrease in growth 
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of the vegetation that surrounds the targeted aquatic system but as discussed below evidence for 

this is mixed. Whether these changes in plant growth are seen as beneficial or adverse will 

depend on the nature of the ecosystem being assessed. 

Information on the effects of changes in nitrogen deposition on forestlands and other 

terrestrial ecosystems is very limited. The multiplicity of factors affecting forests, including 

other potential stressors such as ozone, and limiting factors such as moisture and other nutrients, 

confound assessments of marginal changes in any one stressor or nutrient in forest ecosystems. 

The ISA notes that only a fraction of the deposited nitrogen is taken up by the forests, most of 

the nitrogen is retained in the soils (US EPA, 2008, section 3.3.2.1). In addition, the ISA 

indicates that forest management practices can significantly affect the nitrogen cycling within a 

forest ecosystem, and as such, the response of managed forests to nitrogen deposition will be 

variable depending on the forest management practices employed in a given forest ecosystem 

(US EPA, 2008, Annex C C.6.3). Increases in the availability of nitrogen in nitrogen-limited 

forests via atmospheric deposition could increase forest production over large non-managed 

areas, but the evidence is mixed, with some studies showing increased production and other 

showing little effect on wood production (US EPA, 2008, section 3.3.9). Because leaching of 

nitrate can promote cation losses, which in some cases create nutrient imbalances, slower growth 

and lessened disease and freezing tolerances for forest trees, the net effect of increased N on 

forests in the U.S. is uncertain (US EPA, 2008, section 3.3.9).

The scientific literature has many examples of the deleterious effects caused by excessive 

nitrogen loadings to terrestrial systems. Several studies have set benchmark values for levels of 

N deposition at which scientifically adverse effects are known to occur. Large areas of the 

country appear to be experiencing deposition above these benchmarks. The ISA indicates
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studies that have found that at 3.1 kg N/ha/yr, the community of lichens begins to change from 

acidophytic to tolerant species; at 5.2 kg N/ha/yr, the typical dominance by acidophytic species 

no longer occurs; and at 10.2 kg N/ha/yr, acidophytic lichens are totally lost from the 

community. Additional studies in the Colorado Front Range of the Rocky Mountain National 

Park support these findings. These three values (3.1, 5.2, and 10.2 kg/ha/yr) are one set of 

ecologically meaningful benchmarks for the mixed conifer forest (MCF) of the pacific coast 

regions. Nearly all of the known sensitive communities receive total nitrogen deposition levels 

above the 3.1 N kg/ha/yr ecological benchmark according to the12 km, 2002 CMAQ/NADP 

data, with the exception of the easternmost Sierra Nevadas. The MCFs in the southern portion of 

the Sierra Nevada forests and nearly all MCF communities in the San Bernardino forests receive 

total nitrogen deposition levels above the 5.2 N kg/ha/yr ecological benchmark. 

Coastal Sage Scrub communities are also known to be sensitive to community shifts 

caused by excess nitrogen loadings. Studies have investigated the amount of nitrogen utilized by 

healthy and degraded CSS systems. In healthy stands, the authors estimated that 3.3 kg N/ha/yr

was used for CSS plant growth. It is assumed that 3.3 kg N/ha/yr is near the point where 

nitrogen is no longer limiting in the CSS community and above which level community changes 

occur, including dominance by invasive species and loss of coastal sage scrub. Therefore, this 

amount can be considered an ecological benchmark for the CSS community. The majority of the 

known CSS range is currently receiving deposition in excess of this benchmark. Thus, the REA 

concludes that recent conditions where oxides of nitrogen ambient concentrations are at or below 

the current oxides of nitrogen secondary standards are not adequate to protect against anticipated 

adverse impacts from N nutrient enrichment in sensitive ecosystems. 
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iv. Aquatic nutrient enrichment

The REA aquatic nutrient enrichment case studies focused on coastal estuaries and 

revealed that while current ambient loadings of atmospheric oxides of nitrogen are contributing 

to the overall depositional loading of coastal estuaries, other non-atmospheric sources are 

contributing in far greater amounts in total, although atmospheric contributions are as large as 

some other individual source types. The ability of current data and models to characterize the 

incremental adverse impacts of nitrogen deposition is limited, both by the available ecological 

indicators, and by the inability to attribute specific effects to atmospheric sources of nitrogen. 

The REA case studies used ASSETS EI as the ecological indicator for aquatic nutrient 

enrichment. This index is a six level index characterizing overall eutrophication risk in a water

body. This indictor is not sensitive to changes in nitrogen deposition within a single level of the 

index. In addition, this type of indicator does not reflect the impact of nitrogen deposition in 

conjunction with other sources of nitrogen. 

Based on the above considerations, the REA concludes that the ASSETS EI is not an 

appropriate ecological indicator for estuarine aquatic eutrophication and that additional analysis 

is required to develop an appropriate indicator for determining the appropriate levels of 

protection from N nutrient enrichment effects in estuaries related to deposition of oxides of 

nitrogen. As a result, EPA is unable to make a determination as to the adequacy of the existing 

secondary oxides of nitrogen standard in protecting public welfare from nitrogen nutrient 

enrichment effects in estuarine aquatic ecosystems.

Additionally, nitrogen deposition can alter species composition and cause eutrophication 

in freshwater systems. In the Rocky Mountains, for example, deposition loads of 1.5 to 2 

kg/ha/yr which are well within current ambient levels are known to cause changes in species 
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composition in diatom communities indicating impaired water quality (US EPA, 2008, section 

3.3.5.3). This suggests that the existing secondary standard for oxides of nitrogen does not 

protect such ecosystems and their resulting services from impairment.

v. Other effects

An important consideration in looking at the effects of deposition of oxides of sulfur in 

aquatic ecosystems is the potential for production of MeHg, a neurotoxic contaminant. The 

production of meaningful amounts of MeHg requires the presence of SO4

2-

and mercury, and 

where mercury is present, increased availability of SO4

2-

results in increased production of 

MeHg. There is increasing evidence on the relationship between sulfur deposition and increased 

methylation of mercury in aquatic environments; this effect occurs only where other factors are 

present at levels within a range to allow methylation. The production of MeHg requires the

presence of SO4

2-

and mercury, but the amount of MeHg produced varies with oxygen content, 

temperature, pH and supply of labile organic carbon (US EPA, 2008, section 3.4). In watersheds 

where changes in sulfate deposition did not produce an effect, one or several of those interacting 

factors were not in the range required for meaningful methylation to occur (US EPA, 2008, 

section 3.4). Watersheds with conditions known to be conducive to mercury methylation can be 

found in the northeastern United States and southeastern Canada (US EPA, 2009, section 6). 

With respect to sulfur deposition and mercury methylation, the final ISA determined that 

”[t]he evidence is sufficient to infer a causal relationship between sulfur deposition and increased 

mercury methylation in wetlands and aquatic environments.” However, EPA did not conduct a 

quantitative assessment of the risks associated with increased mercury methylation under current 

conditions. As such, EPA is unable to make a determination as to the adequacy of the existing 
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SO2 secondary standards in protecting against welfare effects associated with increased mercury

methylation.

vi. Summary of Adequacy Considerations

In summary, the PA concludes that currently available scientific evidence and 

assessments clearly call into question the adequacy of the current standards with regard to 

deposition-related effects on sensitive aquatic and terrestrial ecosystems, including acidification 

and nutrient enrichment. Further, the PA recognizes that the elements of the current standards--

indicator, averaging time, level and form – are not ecologically relevant, and are thus not 

appropriate for standards designed to provide such protection. Thus, the PA concludes that 

consideration should be given to establishing a new ecologically relevant multi-pollutant, 

multimedia standard to provide appropriate protection from deposition-related ecological effects 

of oxides of nitrogen and sulfur on sensitive ecosystems, with a focus on protecting against 

adverse effects associated with acidifying deposition in sensitive aquatic ecosystems.

3. CASAC Views

In a letter to the Administrator (Russell and Samet 2011a), the CASAC Oxides of 

Nitrogen and Oxides of Sulfur Panel, with full endorsement of the chartered CASAC, 

unanimously concluded that:

EPA staff has demonstrated through the Integrated Science Assessment (ISA), 

Risk and Exposure Characterization (REA) and the draft PA that ambient NOx 

and SOx can have, and are having, adverse environmental impacts. The Panel 

views that the current NOx and SOx secondary standards should be retained to 

protect against direct adverse impacts to vegetation from exposure to gas phase 

exposures of these two families of air pollutants. Further, the ISA, REA and draft 

PA demonstrate that adverse impacts to aquatic ecosystems are also occurring due 

to deposition of NOx and SOx. Those impacts include acidification and 

undesirable levels of nutrient enrichment in some aquatic ecosystems. The levels 

of the current NOx and SOx secondary NAAQS are not sufficient, nor the forms 

of those standards appropriate, to protect against adverse depositional effects; thus 

a revised NAAQS is warranted.
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In addition, with regard to the joint consideration of both oxides of nitrogen and oxides 

of sulfur as well as the consideration of deposition related effects, CASAC concluded that the PA

had developed a credible methodology for considering such effects. The Panel stated that “the 

Policy Assessment develops a framework for a multi-pollutant, multimedia standard that is 

ecologically relevant and reflects the combined impacts of these two pollutants as they deposit to 

sensitive aquatic ecosystems.”

4. Administrator’s Proposed Conclusions Concerning Adequacy of Current Standard

Based on the above considerations and taking into account CASAC advice, the 

Administrator recognizes that the purpose of the secondary standard is to protect against 

“adverse” effects resulting from exposure to oxides of nitrogen and sulfur, discussed above in 

section II.A. The Administrator also recognizes the need for conclusions as to the adequacy of 

the current standards for both direct and deposition related effects as well as conclusions as to the 

appropriateness and ecological relevance of the current standards. 

In considering what constitutes an ecological effect that is also adverse to the public 

welfare, the Administrator took into account the ISA conclusions regarding the nature and 

strength of the effects evidence, the risk and exposure assessment results, the degree to which the 

associated uncertainties should be considered in interpreting the results, the conclusions 

presented in the PA, and the views of CASAC and members of the public. On these bases, the 

Administrator concludes that the current secondary standards are adequate to protect against 

direct phytotoxic effects on vegetation. Thus, the Administrator proposes to retain the current 

secondary standard for oxides of nitrogen at 53 ppb,4annual average concentration, measured in 

 

4 The annual secondary standard for oxides of nitrogen is being specified in units of ppb to conform to the current 

version of the annual primary standard, as specified in the final rule for the most recent review of the NO2 primary 

NAAQS (75 FR 6531; February 9, 2010).
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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. 

With regard to deposition-related effects, the Administrator has first to consider the 

appropriateness of the structure of the current standards to address ecological effects of concern. 

Based on the evidence as well as considering the advice given by CASAC on this matter, the 

Administrator concludes that the elements of the current standards are not ecologically relevant 

and thus are not appropriate to provide protection of ecosystems. On the subject of adequacy of 

protection with regard to deposition-related effects, the Administrator considered the full nature 

of ecological effects related to the deposition of ambient oxides of nitrogen and sulfur into 

sensitive ecosystems across the country. Her conclusions are based on the evidence presented in 

the ISA with regard to acidification and nutrient enrichment effects, the findings of the REA with 

regard to scope and severity of the current and likely future effects of deposition, the synthesis of 

both the scientific evidence and risk and exposure results in the PA as to the adequacy of the 

current standards, and the advice of both CASAC and the public. After such consideration, the 

Administrator concludes that current levels of oxides of nitrogen and sulfur are sufficient to 

cause acidification of both aquatic and terrestrial ecosystems, nutrient enrichment of terrestrial 

ecosystems and contribute to nutrient enrichment effects in estuaries that could be considered 

adverse, and the current secondary standards do not provide adequate protection from such 

effects. 

Having reached these conclusions, the Administrator determines that it is appropriate to 

consider alternative standards that are ecologically relevant. These considerations support the 

conclusion that the current secondary standards is neither appropriate nor adequate to protect 
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against deposition related effects. The Administrator’s consideration of such alternative 

standards is discussed below in Section III.

III. Rationale for Proposed Decision on Alternative Multi-pollutant Approach to

Secondary Standards for Aquatic Acidification

Having reached the conclusion that the current NO2 and SO2 secondary standards are not 

adequate to provide appropriate protection against deposition-related effects associated with 

oxides of nitrogen and sulfur, the Administrator then considered what new multi-pollutant 

standard might be appropriate, at this time, to address such effects on public welfare. The 

Administrator recognizes that the inherently complex and variable linkages between ambient 

concentrations of nitrogen and sulfur oxides, the related deposited forms of nitrogen and sulfur, 

and the ecological responses that are associated with public welfare effects call for consideration 

of an ecologically relevant design of a standard that reflects these linkages. The Administrator 

also recognizes that characterization of such complex and variable linkages will necessarily 

require consideration of information and analyses that have important limitations and 

uncertainties.

Despite its complexity, an ecologically relevant multi-pollutant standard to address 

deposition-related effects could still appropriately be defined in terms of the same basic elements 

that are used to define any NAAQS – indicator, form, averaging time, and level. The form 

would incorporate additional structural elements that reflect relevant multi-pollutant and 

multimedia attributes. These structural elements include the use of an ecological indicator, tied 

to the ecological effect we are focused on, and other elements that account for ecologically

relevant factors other than ambient air concentrations. All of these elements would be needed to 

enable a linkage from ambient air indicators to the ecological indicator to define an ecologically 
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relevant standard. As a result, such a standard would necessarily be more complex than the 

NAAQS that have been set historically to address effects associated with ambient concentrations 

of a single pollutant.

More specifically, the Administrator considered an ecologically relevant multi-pollutant

standard to address effects associated with acidifying deposition related to ambient 

concentrations of oxides of nitrogen and sulfur in sensitive aquatic ecosystems. This focus is 

consistent with the information presented in the ISA, REA, and PA, which highlighted the 

sufficiency of the quantity and quality of the available evidence and assessments associated with 

aquatic acidification relative to the information and assessments available for other depositionrelated effects, including terrestrial acidification and aquatic and terrestrial nutrient enrichment. 

Based on its review of these documents, CASAC agreed that aquatic acidification should be the 

focus for developing a new multi-pollutant standard in this review. In reaching conclusions 

about an air quality standard designed to address deposition-related aquatic acidification effects, 

the Administrator also recognizes that such a standard may also provide some degree of 

protection against other deposition-related effects.

As discussed in chapter 7 of the PA, the development of a new multi-pollutant standard to 

address deposition-related aquatic acidification effects recognizes the need for consideration of a 

nationally applicable standard for protection against adverse effects of aquatic acidification on

public welfare, while recognizing the complex and heterogeneous interactions between ambient 

air concentrations of nitrogen and sulfur oxides, the related deposition of nitrogen and sulfur, and 

associated ecological responses. The development of such a standard also needs to take into 

account the limitations and uncertainties in the available information and analyses upon which 

characterization of such interactions are based. The approach used in the PA also recognizes that 
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while such a standard would be national in scope and coverage, the effects to public welfare 

from aquatic acidification will not occur to the same extent in all locations in the U.S., given the 

inherent variability of the responses of aquatic systems to the effects of acidifying deposition.

As discussed above in section II, many locations in the U.S. are naturally protected 

against acid deposition due to underlying geological conditions. Likewise, some locations in the 

U.S., including lands managed for commercial agriculture and forestry, are not likely to be 

negatively impacted by current levels of nitrogen and sulfur deposition. As a result, while a new 

ecologically relevant secondary standard would apply everywhere, it would be structured to 

account for differences in the sensitivity of ecosystems across the country. This would allow for 

appropriate protection of sensitive aquatic ecosystems, which are relatively pristine and wild and 

generally in rural areas, and the services provided by such sensitive ecosystems, without 

requiring more protection than is needed elsewhere. 

As discussed below, the multi-pollutant standard developed in the PA would employ (1) 

total reactive oxidized nitrogen (NOy) and SOx as the atmospheric ambient air indicators; (2) a 

form that takes into account variable factors, such as atmospheric and ecosystem conditions that 

modify the amounts of deposited nitrogen and sulfur; the distinction between oxidized and 

reduced forms of nitrogen; effects of deposited nitrogen and sulfur on aquatic ecosystems in 

terms of the ecological indicator ANC; and the representativeness of water bodies within a 

defined spatial area; (3) a multi-year averaging time, and (4) a standard level defined in terms of 

a single, national target ANC value that, in the context of the above form, identifies the levels of 

concentrations of NOy and SOx in the ambient air that would meet the standard. The form of 

such a standard has been defined by an index, AAI, which reflects the relationship between 
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ambient concentrations of NOy and SOx and aquatic acidification effects that result from nitrogen 

and sulfur deposition related to these ambient concentrations.

In presenting the considerations associated with such an air quality standard to address 

deposition-related aquatic acidification effects, the following sections focus on each element of 

the standard, including indicator (section III.A), form (section III.B), averaging time (section 

III.C), and level (section III.D). Alternative combinations of levels and forms are discussed in 

section III.E. Considerations related to important uncertainties inherent in such an approach are 

discussed in section III.F. Advice from CASAC on such a new standard is presented in section 

III.G. The Administrator’s proposed decisions on such a new standard are presented in section 

III.H.

A. Ambient Air Indicators

In considering alternative ambient air indicators, the PA primarily focuses on the 

important attribute of association. Association in a broad sense refers to how well an ambient air 

indicator relates to the ecological effects of interest by virtue of both the framework that links the 

ambient indicator and effects and the empirical evidence that quantifies the linkages. The PA

also considers how measurable or quantifiable an indicator is to enable its use as an effective 

indicator of relevant ambient air concentrations.

As discussed above in section II.C, the PA concludes that indicators other than NO2 and 

SO2 should be considered as the appropriate indicators of oxides of nitrogen and sulfur in the 

ambient air for protection against the acidification effects associated with deposition ofthe 

associated nitrogen and sulfur. This conclusion is based on the recognition that all forms of 

nitrogen and sulfur in the ambient air contribute to deposition and resulting acidification, and as 

such, NO2 and SO2 are incomplete indicators. In principle, the ambient indicators should 
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represent the species that are associated with oxides of nitrogen and sulfur in the ambient air and 

can contribute acidifying deposition. This includes both the species of oxides of nitrogen and 

sulfur that are directly emitted as well as species transformed in the atmosphere from oxides of 

nitrogen and sulfur that retain the nitrogen and sulfur atoms from directly emitted oxides of 

nitrogen and sulfur. All of these compounds are associated with oxides of nitrogen and sulfur in 

the ambient air and can contribute to acidifying deposition. 

The PA focuses in particular on the various compounds with nitrogen or sulfur atoms that 

are associated with oxides of nitrogen and sulfur, because the acidifying potential is specific to 

nitrogen and sulfur, and not other atoms (e.g., H, C, O) whether derived from the original source 

of oxides of nitrogen and sulfur emissions or from atmospheric transformations. For example, 

the acidifying potential of each molecule of NO2, NO, HNO3 or PAN is identical, as is the 

potential for each molecule of SO2 or ion of particulate sulfate, p-SO4. Each atom of sulfur 

affords twice the acidifying potential of each atom of nitrogen.

1. Oxides of Sulfur

As discussed in the PA (US EPA, 2011, section 7.1.1), oxides of sulfur include the gases 

sulfur monoxide (SO), SO2, sulfur trioxide (SO3), disulfur monoxide (S2O), and particulatephase sulfur compounds (referred to as SO4) that result from gas-phase sulfur oxides interacting 

with particles. However, the sum of SO2 and SO4 does represent virtually the entire ambient air 

mass of sulfur that contributes to acidification. In addition to accounting for virtually all the 

potential for acidification from oxidized sulfur in the ambient air, there are reliable methods to 

monitor the concentrations of SO2 and particulate SO4. In addition, much of the data used to 

develop the technical basis for the standard developed in the PA is based on monitoring or 
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modeling of these species.5 The PA concludes that the sum of SO2 and SO4, referred to as SOx, 

are appropriate ambient air indicators of oxides of sulfur because they represent virtually all of 

the acidification potential of ambient air oxides of sulfur and there are reliable methods suitable 

for measuring SO2 and SO4.

2. Oxides of Nitrogen

As discussed in the PA (US EPA, 2011,section 7.1.2), NOy, as defined in chapter 2 of the 

PA, incorporates basically all of the oxidized nitrogen species that have acidifying potential and 

as such, NOy should be considered as an appropriate indicator for oxides of nitrogen. Total 

reactive oxidized nitrogen is an aggregate measure of NO and NO2 and all of the reactive 

oxidized products of NO and NO2. That is, NOy is a group of nitrogen compounds in which all 

of the compounds are either an oxide of nitrogen or compounds in which the nitrogen atoms 

came from oxides of nitrogen. Total reactive oxidized nitrogen is especially relevant as an 

ambient indicator for acidification in that it both relates to the oxides of nitrogen in the ambient 

air and also represents the acidification potential of all oxidized nitrogen species in the ambient 

air, whether an oxide of nitrogen or derived from oxides of nitrogen. 

There are currently available reliable methods of measuring aggregate NOy. The term 

“aggregate” measure means that the NOy, as measured, is not based on measuring each 

individual species of NOy and calculating an NOy value by summing the individual species. 

Rather, as described in chapter 2 of the PA, current measurement techniques process all of the 

individual NOy species to produce a single aggregate measure of all of the nitrogen atoms 

associated with any NOy species. Consequently, the NOy measurement effectively provides the 

 

5 As discussed in chapter 2 of the PA, SO2 and particulate SO4 are routinely measured in ambient air monitoring 

networks, although only the Clean Air Status and Trends Network (CASTNET) filter packs do not intentionally 

exclude particle size fractions. The CMAQ treatment of SOx is the simple addition of both species, which are 

treated explicitly in the model formulation. All particle size fractions are included in the CMAQ SOx estimates.
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sum of all individual species, but the identity of the individual species is lost. As discussed 

above, the accounting for the individual nitrogen atoms is an accounting of the ambient air 

acidification potential of oxides of nitrogen and their transformation products and therefore the 

most relevant ambient indicator for aquatic acidification effects associated with oxides of 

nitrogen. 

This loss of the information on individual species motivated consideration of alternative 

or more narrowly defined indicators for oxides of nitrogen in the PA. Consideration of a subset 

of NOy species was based on the following reasoning. First, the actual dry deposition of nitrogen 

is determined on an individual species basis by multiplying the species concentration times a 

species-specific deposition velocity and then summed to develop an estimate of total dry

deposition. Consequently, the use of individual ambient species has the potential to be more 

consistent with the underlying science of deposition and, therefore, has the potential to allow for 

a more rigorous evaluation of dry deposition with specialized field studies. In addition, there has 

been a suggestion of focusing only on the most quickly depositing NOyspecies, such as HNO3, as 

contributions from other NOy species such as NO2 may be negligible. These alternative 

indicators are discussed below.

The PA considers the relative merits of using each individual NOy species as part of a 

group of indicators. In so doing, it was first noted that dry deposition of NOy is treated as the 

sum of the deposition of each individual species in advanced process-based air quality models 

like CMAQ, as described in chapter 2 of the PA. Conceptually one could extend this processbased approach by using all NOy species individually as separate indicators for oxides of 

nitrogen and requiring, for example, measurements of each of the species, including the 

dominant species of HNO3, particulate nitrate (p-NO3), true NO2, NO and PAN. The potential 
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attraction of using individual species would be the reliance on actual deposition velocities. This 

could have more physical meaning in comparison to a constructed model of aggregate deposition

of NOy, which is difficult to evaluate with observations because of the assimilation of many

species with disparate deposition behavior. The PA notes that the major drawback of using 

individual NOy species as the indicators is the lack of reliable measurement techniques, 

especially for PAN and NO2 in rural locations, which renders the use of virtually any individual 

NOy species, except for NO and perhaps p-NO3, as functionally inadequate from a measurement 

perspective.

The PA next considered the relative merits of using a subset of NOy species as the 

indicators for oxides of nitrogen, as was discussed above for oxides of sulfur. To the extent that 

certain species provide relatively minor contributions to total NOy deposition, it may be 

appropriate to consider excluding them as part of the indicator. As discussed in chapter 2 of the 

PA, each nitrogen species within the array of NOy species has species-specific dry deposition 

velocities. For example, the deposition velocity of HNO3 is much greater than the velocity for 

NO2 and, consequently, for a similar ambient air concentration, HNO3 contributes more 

deposition of acidifying nitrogen relative to NO2. In transitioning from source-oriented urban 

locations to rural environments, the ratio of the concentrations of HNO3 and PAN to NO2

increases.

Based on the reasoning that a larger fraction of the deposited NOy is accounted for by

total nitrate (the sum of HNO3 and p-NO3), a surrogate for the more rapidly depositing fraction 

of NOy, combined with the availability of reliable total nitrate measurements through the 

CASTNET, the PA considered using total nitrate as the indicator for oxides of nitrogen (US 

EPA, 2011, appendix E). Nitrate would be expected to correlate well with total reactive oxidized 
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nitrogen deposition relative to NOy (US EPA, 2011, chapter 2) despite the inherent noise 

associated with variable contributions of low deposition velocity species (e.g., NO2) that may

have relatively high ambient concentrations. However, modeling simulations suggest that NOy

may be a more robust indicator, relative to HNO3, in terms of relating absolute changes in 

ambient air concentrations to changes in nitrogen deposition driven by changes in ambient 

concentrations of oxides of nitrogen (US EPA, 2011, Figure 2-32).

Based on the above considerations, the PA concludes that NOy should be considered as 

the appropriate ambient indicator for oxides of nitrogen based on its direct relationship to oxides 

of nitrogen in the ambient air and its direct relationship to deposition associated with aquatic 

acidification. Because NOy represents all of the potentially acidifying oxidized nitrogen species 

in the ambient air, it is appropriately associated with the deposition of potentially acidifying 

compounds associated with oxides of nitrogen in the ambient air. In addition, there are reliable 

methods available to measure NOy. Measurement of each individual species of NOy, or the 

measurement of only a subset of species of NOy, is less appropriate because there are not reliable 

measurements methods available to measure all of the individual species of NOy and a subset of 

species would fail to account for significant portions of the oxidized reactive nitrogen that relate 

to acidification.6

B. Form

Based on the evidence of the aquatic acidification effects caused by the deposition of 

NOy and SOx, the PA (US EPA, 2011,section 7.2) presents the development of a new form that is 

ecologically relevant for addressing such effects. The conceptual design for the form of such a 

standard includes three main components: an ecological indicator, deposition metrics that relate 

 

6 The PA also notes that NOy is a useful measurement for model evaluation purposes, which is especially important, 

recognizing the unique role that CMAQ plays in the development of this standard, as described below in section 

III.B.
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to the ecological indicator, and a function that relates ambient air indicators to deposition 

metrics. Collectively, these three components link the ecological indicator to ambient air 

indicators, as illustrated above in Fig II-1.

The simplified flow diagram in Figure II-1 compresses the various atmospheric, 

biological, and geochemical processes associated with acidifying deposition to aquatic

ecosystems into a simplified conceptual picture. The ecological indicator (left box) is related to 

atmospheric deposition through biogeochemical ecosystem models (middle box), which 

associate a target deposition load to a target ecological indicator. Once a target deposition is 

established, associated allowable air concentrations are determined (right box) through the 

relationships between concentration and deposition that are embodied in air quality models such 

as CMAQ. The following discussion describes the development and rationale for each of these 

components, as well as the integration of these components into the full expression of the form 

of the standard using the concept of a national AAI that represents a target ANC level as a 

function of ambient air concentrations. Spatial aggregation issues associated with defining each 

of the terms of this index are also addressed below.

The AAI is designed to be an ecologically relevant form of the standard that determines 

the levels of NOy and SOx in the ambient air that would achieve a target ANC limit for the U.S. 

The intent of the AAI is to weight atmospheric concentrations of oxides of nitrogen and sulfur by 

their propensity to contribute to acidification through deposition, given the fundamental 

acidifying potential of each pollutant, and to take into account the ecological factors that govern 

acid sensitivity in different ecosystems. The index also accounts for the contribution of reduced 

nitrogen to acidification. Thus, the AAI encompasses those attributes of specific relevance to 
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protecting ecosystems from the acidifying potential of ambient air concentrations of NOy and 

SOx.

1. Ecological Indicator

In considering alternative ecological indicators, the PA again primarily focuses on the 

attribute of association. In the case of an ecological indicator for aquatic acidification, 

association refers to the relationship between the indicator and adverse effects as discussed in 

section II. Because of the conceptual structure of the form of an AAI-based standard (Figure III1), this particular ecological indicator must also link up in a meaningful and quantifiable manner 

with acidifying atmospheric deposition. In effect, the ecological indicator for aquatic 

acidification is the bridge between biological impairment and deposition of NOyand SOx. 

This section presents the rationale in the PA for selecting ANC as the appropriate 

ecological indicator for consideration. Recognizing that ANC is not itself the causative or toxic 

agent for adverse aquatic acidification effects, the rationale for using ANC as the relevant 

ecological indicator is based on the following:

(1) The ANC is directly associated with the causative agents, pH and dissolved Al, both 

through empirical evidence and mechanistic relationships;

(2) Empirical evidence shows very clear and strong relationships between adverse effects 

and ANC;

(3) The ANC is a more reliable indicator from a modeling perspective, allowing use of a 

body of studies and technical analyses related to ANC and acidification to inform the 

development of the standard; and
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(4) The ANC literally embodies the concept of acidification as posed by the basic principles 

of acid base chemistry and the measurement method used to estimate ANC and, 

therefore, serves as a direct index to protect against acidification. 

Ecological indicators of acidification in aquatic ecosystems can be chemical or biological 

components of the ecosystem that are altered by the acidifying effects of nitrogen and sulfur 

deposition. A desirable ecological indicator for aquatic acidification is one that is measurable or 

estimable, linked causally to deposition of nitrogen and sulfur, and linked causally, either 

directly or indirectly to ecological effects known or anticipated to adversely affect public 

welfare.

As summarized in chapter 2 of the PA, atmospheric deposition of NOy and SOx causes 

aquatic acidification through the input of strong acid anions (e.g., NO3

-

and SO4

2-

) that ultimately

shifts the water chemistry equilibrium toward increased hydrogen ion levels (or decreased pH). 

The anions are deposited either directly to the aquatic ecosystem or indirectly via transformation 

through soil nitrification processes and subsequent drainage from terrestrial ecosystems. In other 

words, when these anions are mobilized in the terrestrial soil, they can leach into adjacent water 

bodies. Aquatic acidification is indicated by changes in the surface water chemistry of 

ecosystems. In turn, the alteration of surface water chemistry has been linked to negative effects 

on the biotic integrity of freshwater ecosystems. There is a suite of chemical indicators that 

could be used to assess the effects of acidifying deposition on lake or stream acid-base 

chemistry. These indicators include ANC; alkalinity (ALK); base neutralizing capacity, 

commonly referred to as acidity (ACY); surface water pH; concentrations of trivalent aluminum, 

Al+3; and concentrations of major anions (SO4

2-

, NO3

-

), cations (Ca2+, Mg+2, K+), or sums of 

cations or anions.
Page 115 of 238

The ANC and ALK are very similar quantities and are used interchangeably in the 

literature and for some of the analyses presented in this document. Both ANC and ALK are 

defined as the amount of strong acid required to reach a specified equivalence point. For acidbase solutions, an equivalence point can be thought of as the point at which the addition of strong 

acids (i.e., titration) is no longer neutralized by the solution. This explains the term acid 

neutralizing capacity, or ANC, as ANC relates directly to the capacity of a system to neutralize 

acids. The differences between ANC and ALK are based on operational definitions and subject 

to various interpretations. The ANC is preferred over ALK as the body of scientific evidence 

has focused on ANC and effects relationships. The ALK is more widely associated with more 

general characterizations of water quality such as the relative hardness of water associated with 

carbonates.

Indictors such as the concentrations of specific anions, cations, or their groupings, while 

relevant to acidification processes, are not robust acidification indicators as it is the relative 

balance of cations and anions that is more directly associated with acidification. That balance is 

captured by ANC and ALK. Acidity, ACY, is the converse of ANC and indicates how much 

strong base it takes to reach an equivalence point. Because ACY is not used in most ecosystem 

assessments, the body of information relating ACY to effects is too limited to serve as a basis for 

an appropriate ecological indicator. Aluminum and other metals are causative toxic agents that 

directly impair biological functions. However, Al, or metals in general, have high variability in 

concentrations that can be linked to effects, often at extremely low levels which in some cases 

approach detectability limits, exhibit rapid transient responses, and are often confounded by the 

presence of other toxic metals. These concerns limit the use of metals as reliable and measurable 

ecological indicators. Hydrogen ion (H+) concentrations, using their negative logarithmic values, 
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or pH, are well correlated with adverse effects, as discussed above in section II.A, and determine 

the solubility of metals such as aluminum. However, pH is not a preferred acidification indicator 

due to its highly transient nature and other concerns, as discussed below. 

Having reasoned that ANC is a preferred indicator to ALK, ACY, individual metals or 

groupings of ions, the PA considers the relative merits of ANC compared to pH, which is a well 

recognized indicator of acidity and a more direct causative agent with regard to adverse effects. 

First, the linkage between ANC and pH is considered in recognition of the causative association

between pH and effects.

The ANC is not the direct causative toxic agent impacting aquatic species diversity. The 

scientific literature generally emphasizes the links between pH and adverse effects as described 

above in section II.A. It is important, therefore, to consider the extent to which ANC and pH are 

well related from a mechanistic perspective as well as through empirical evidence. The ANC 

and pH are co-dependent on each other based on the requirement that all solutions are 

electrically neutral, meaning that any solution must satisfy the condition that all negatively

charged species must be balanced by all positively charged species. The ANC is defined as the 

difference between strong anions and cations (US EPA, 2011, equation 7-13).

While the chemistry can be complex, the co-dependency between ANC and pH is 

explained by recognizing that positively charged hydrogen, H+, is incorporated in the charge 

balance relationships related to the overall solution chemistry which also defines ANC. The 

positive, directional co-dependency (i.e., ANC and pH increase together) is further explained in 

concept as ANC reflects how much strong acid (i.e., how much hydrogen ion) it takes to titrate to 

an equivalence point. Strong observed correlations between pH and ANC as described in the PA

support these mechanistic relationships. 
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As discussed above in section II.A, there are well established examples of ANC 

correlating strongly with a variety of ecological effects which are summarized in the PA (US 

EPA, 2011, Table 3-1). Because pH and ANC are well correlated and linearly dependent over 

the pH ranges (4.5-6) where adverse ecological effects are observed, evidence of clear 

associations exist between ANC and adverse ecological effects as described in the PA. In large 

measure, this dependence between pH and ANC and the relationship of both pH and ANC to 

effects, speak directly to the appropriateness of ANC with respect to its use as an ecological 

indicator.

Thus, there is a clear association between ANC and ecological effects, although there is a 

more direct causal relationship between pH and ecological effects. Nonetheless, ANC is 

preferred as an ecological indicator based on its superior ability to provide a linkage with 

deposition in a meaningful and quantifiable manner, a role that is served far more effectively by

ANC than by pH. While both ANC and pH are clearly associated with the effects of concern, 

ANC is superior in linking these effects to deposition. 

The PA notes that the basis for this conclusion is that acidifying atmospheric deposition 

of nitrogen and sulfur is a direct input of potential acidity (ACY), or, in terms of ANC, such 

deposition is relevant to the major anions that reduce the capacity of a water body to neutralize 

acidity. Consequently, there is a well defined linear relationship between potential acidifying 

deposition and ANC. This ANC-deposition relationship facilitates the linkage between 

ecosystem models that calculate an ecological indicator and the atmospheric deposition of NOy

and SOx. On the other hand, there is no direct linear relationship between deposition and pH. 

While acid inputs from deposition lower pH, the relationship can be extremely nonlinear and 

there is no direct connection from a modeling or mass balance perspective between the amount 
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of deposition entering a system and pH. The term “mass balance” underlies the basic 

formulation of any physical modeling construct, for atmospheric or aquatic systems, and refers to 

the accounting of the flow of mass into a system, the transformation to other forms, and the loss 

due to flow out of a system and other removal processes. The ANC is a conserved property. 

This means that ANC in a water body can be accounted for by knowledge of how much ANC 

initially exists, how much flows in and is deposited, and how much flows out. In contrast, 

hydrogen ion concentration in the water, the basis for pH, is not a conserved property as its 

concentration is affected by several factors such as temperature, atmospheric pressure, mixing 

conditions of a water body, and the levels of several other chemical species in the system. The 

disadvantage of pH lacking conservative properties is that there is a very complex connection 

between changes in ambient air concentrations of NOy and SOx and pH. 

The discussion of basic water chemistry of natural systems in chapter 2 of the PA

provides further details on why pH is not a conserved quantity and is subject to rapid transient 

response behavior that makes it difficult to use as a reliable and functional ecological indicator. 

The observed pH-to-ANC relationship (US EPA, 2011, figure 7-2) partially explains the concern 

with pH responding too abruptly. In the region where pH ranges roughly from 4.5 to 6 and is of 

greatest relevance to effects (US EPA, 2011, figure 7-4), there clearly is more sensitivity of pH 

to changes in ANC in the ANC range from approximately 0 to 50 μeq/L. A focus on this part of 

the ANC-to-pH relationship shows that ANC associates well with pH in a fairly linear manner. 

However, the pH range from 4.5 to 6 also includes one of the very steepest parts of the slope 

relating pH as a function of ANC, where ANC ranges down below 0 μeq/L, which is subject to 

very rapid change in ANC, or deposition inputs. This part of the relationship coincides with 

reduced levels of ANC and hence with reduced ability to neutralize acids and moderate pH 
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fluctuations. This response behavior can be extended to considering how pH would change in 

response to deposition, or ambient concentrations, of NOy and SOx, which can be viewed as 

“ANC-like” inputs. 

In summary, because ANC clearly links both to biological effects of aquatic acidification 

as well as to acidifying inputs of NOy and SOx deposition, the PA concludes that ANC is an 

appropriate ecological indicator for relating adverse aquatic ecosystem effects to acidifying 

atmospheric deposition of SOx and NOy, and is preferred to other potential indicators. In 

reaching this conclusion, the PA notes that in its review of the first draft PA, CASAC concluded 

that “information on levels of ANC protective to fish and other aquatic biota has been well 

developed and presents probably the lowest level of uncertainty in the entire methodology” 

(Russell and Samet, 2010a). In its more recent review of the second draft PA, CASAC agreed 

“that acid neutralizing capacity is an appropriate ecological measure for reflecting the effects of 

aquatic acidification” (Russell and Samet, 2010b; p. 4). 

2. Linking ANC to Deposition

There is evidence to support a quantified relationship between deposition of nitrogen and 

sulfur and ANC. This relationship was analyzed in the REA for two case study areas, the 

Adirondack and Shenandoah Mountains, based on time-series modeling and observed trends. 

In the REA analysis, long-term trends in surface water nitrate, sulfate and ANC were modeled 

using MAGIC for the two case study areas. These data were used to compare recent surface 

water conditions in 2006 with preindustrial conditions (i.e. preacidification 1860). The results 

showed a marked increase in the number of acid impacted lakes, characterized as a decrease in 

ANC levels, since the onset of anthropogenic nitrogen and sulfur deposition, as discussed in 

chapter 2 of the PA.
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In the REA, more recent trends in ANC, over the period from 1990 to 2006, were 

assessed using monitoring data collected at the two case study areas. In both case study areas, 

nitrate and sulfate deposition decreased over this time period. In the Adirondack Mountains, this 

corresponded to a decreased concentration of nitrate and sulfate in the surface waters and an 

increase in ANC (US EPA, 2009, section 4.2.4.2). In the Shenandoah Mountains, there was a 

slight decrease in nitrate and sulfate concentration in surface waters corresponding to modest 

increase in ANC from 50 μeq/L in 1990 to 67 μeq/L in 2006 (US EPA, 2009, section 

4.2.4.3,Appendix 4, and section 3.4).

In the REA, the quantified relationship between deposition and ANC was investigated 

using ecosystem acidification models, also referred to as acid balance models or critical loads 

models (US EPA, 2011, section 2 and US EPA, 2009, section 4 and Appendix 4). These models 

quantify the relationship between deposition of nitrogen and sulfur and the resulting ANC in 

surface waters based on an ecosystem’s inherent generation of ANC and ability to neutralize 

nitrogen deposition through biological and physical processes. A critical load is defined as the 

amount of acidifying atmospheric deposition of nitrogen and sulfur beyond which a target ANC 

is not reached. Relatively high critical load values imply that an ecosystem can accommodate 

greater deposition levels than lower critical loads for a specific target ANC level. Ecosystem 

models that calculate critical loads form the basis for linking deposition to ANC. 

As discussed in chapter 2 of the PA, both dynamic and steady state models calculate 

ANC as a function of ecosystem attributes and atmospheric nitrogen and sulfur deposition, and 

can be used to calculate critical loads. Steady state models are time invariant and reflect the 

long term consequences associated with an ecosystem reaching equilibrium under a constant 

level of atmospheric deposition. Dynamic models are time variant and take into account the 
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time dependencies inherent in ecosystem hydrology, soil and biological processes. Dynamic 

models like MAGIC can provide the time series response of ANC to deposition whereas steady

state models provide a single ANC relationship to any fixed deposition level. Dynamic models 

naturally are more complex than steady state models as they attempt to capture as much of the 

fundamental biogeochemical processes as practicable, whereas steady state models depend on far 

greater parameterization and generalization of processes that is afforded, somewhat, by not 

having to accounting for temporal variability. 

The PA notes that steady state models are capable of addressing the question of what 

does it take to reach and sustain a specific level of ANC. Dynamic models are also capable of 

addressing that question, but can also address the question of how long it takes to achieve that 

result. Dynamic models afford the ability for more comprehensive treatment of a variety of 

processes throughout the surface, soil and bedrock layers within an ecosystem. For example, 

steady state models treat sulfate as a mobile anion throughout the system, meaning that the 

sulfate that is deposited to a watershed enters the water column and is not influenced by soil 

adsorption or cation exchange. Dynamic models can incorporate these time variant processes. 

The use of a steady state model treating sulfate as totally mobile does not necessarily conflict 

with the possibility of sulfate acting as a less than mobile ion at certain times. The steady state 

assumption is premised on the long term behavior of sulfate which can undergo periods of net 

adsorption followed by periods of net desorption which can balance out over time. The PA

recognizes that as the richness of the available data increases, in terms of parameters and spatial 

resolution, the incorporation of dynamic modeling approaches in the standard setting process 

should become more feasible. In determining an appropriate modeling approach for the 

development of a NAAQS in this review, the PA considers both the relevance of the question 
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addressed as well as the ability to perform modeling that provides relevant information for 

geographic areas across the country. 

Dynamic models require a large amount of catchment level-specific data relative to 

steady state models. Because of the time invariant nature of steady state models, the data 

requirements that integrate across a broad spectrum of ecosystem processes is achievable and 

available now at the national level. Water quality data to support steady state models currently 

exist for developing a national data base for modeling nearly 10,000 catchments in the 

contiguous U.S. In contrast, the data needs to support dynamic models for national-scale 

analyses simply are not available at this time. Further, the information provided by steady state 

modeling would be sufficient to develop and analyze alternative NAAQS and the kind of 

protection they would afford. While it would be of interest to also obtain information about how 

much time it would take for a target ANC level to be achieved, the absence of such information 

does not preclude developing and evaluating alternative NAAQS using the AAI structure. Based 

on the above considerations, the PA concludes that at this time steady state critical load modeling 

is an appropriate tool for linking long-term ANC levels to atmospheric deposition of nitrogen 

and sulfur for development of an AAI that has national applicability.

A steady state model is used to define the critical load, which is the amount of 

atmospheric deposition of nitrogen (N) and sulfur (S) beyond which a target ANC is not 

achieved and sustained.7 It is expressed as:

CLANClim(N + S) = ([BC]0

*

- [ANClim])Q + Neco (III-1)

 

7

 This section discusses the linkages between deposition of nitrogen and sulfur and ANC. Section III.B.3 then 

discusses the linkages between atmospheric concentrations of NOy and SOx and deposition of nitrogen and sulfur.
Page 123 of 238

Where:

CLANClim(N + S) is the critical load of deposition, with units of equivalent charge/(area-time);

[BC]0

*

is the natural contribution of base cations from weathering, soil processes and 

preindustrial deposition, with units of equivalent charge/volume;

[ANClim] is the target ANC value, with units of equivalent charge/volume; 

Q is the catchment level runoff rate governed by water mass balance and dominated by

precipitation, with units of distance/time; and

Neco is the amount of nitrogen deposition that is effectively neutralized by a variety of biological 

(e.g., nutrient uptake) and physical processes, with units of equivalent charge/ (area-time).

Equation III-1 is a modified expression that adopts the basic formulation of the steady

state models that are described in chapter 2 of the PA. More detailed discussion of the rationale, 

assumptions and derivation of equation III-1, as well as all of the equations in this section, are 

included in Appendix B of the PA. The equation simply reflects the amount of deposition of 

nitrogen and sulfur from the atmosphere, CLANClim(N + S), that is associated with a sustainable 

long-term ANC target, [ANClim], given the capacity of the natural system to generate ANC, 

[BC]0

*

, and the capacity of the natural system to neutralize nitrogen deposition, Neco. This 

expression of critical load is valid when nitrogen deposition is greater than Neco.

8

 The runoff 

rate, Q, allows for balancing mass in the two environmental mediums – atmosphere and 

catchment. This critical load expression can be focused on a single water system or more

broadly. To extend applicability of the critical load expression (equation III-1) from the 

catchment level to broader spatial areas, the terms Qr and CLr, are used, which are the runoff rate 

 

8 Because Neco is only relevant to nitrogen deposition, in rare cases where Neco is greater than the total nitrogen 

deposition, the critical load would be defined only in terms of acidifying deposition of sulfur and the Neco term in 

equation III-1 would be set to zero.
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and critical load, respectively, of the region over which all the atmospheric terms in the equation 

are defined. 

In considering the contributions of SOx or NOy species to acidification, it is useful to 

think of every depositing nitrogen atom as supplying one equivalent charge unit and every sulfur 

atom as depositing two charge units. The PA uses equivalent charge per volume as a 

normalizing tool in place of the more familiar metrics such as mass or moles per volume. This 

allows for a clearer explanation of many of the relationships between atmospheric and ecosystem 

processes that incorporate mass and volume unit conventions somewhat specific to the 

environmental media of concern (e.g., m3for air and liter for liquid water). Equivalent charge 

reflects the chemistry equilibrium fundamentals that assume electroneutrality, or balancing 

charge where the sum of cations always equals the sum of anions. 

As presented above, the terms S and N in the CLANClim (N + S) term broadly represent all 

species of sulfur or nitrogen that can contribute to acidifying deposition. This follows 

conventions used in the scientific literature that addresses critical loads, and it reflects all 

possible acidifying contributions from any sulfur or nitrogen species. For all practical purposes, 

S reflects SOx as described above, the sum of sulfur dioxide gas and particulate sulfate. 

However, N in equation III-1 includes both oxidized forms, consistent with the ambient 

indicator, NOy, in addition to the reduced nitrogen species, ammonia and ammonium ion, 

referred to as NHx. The NHx is included in the critical load formulation because it contributes to 

potentially acidifying nitrogen deposition. Consequently, from a mass balance or modeling 

perspective, the form of the standard needs to account for NHx, as described below.

3. Linking Deposition to Ambient Air Indicators
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The last major component of the form illustrated in Figure III-1 addresses the linkage 

between deposition of nitrogen and sulfur and concentrations of the ambient air indicators, NOy

and SOx. To link ambient air concentrations with deposition, the PA defines a transference 

ratio, T, as the ratio of total wet and dry deposition to ambient concentration, consistent with the 

area and time period over which the standard is defined. To express deposition of NOy and SOx

in terms of NOy and SOx ambient concentrations, two transference ratios were defined, where 

TSOx equals the ratio of the combined dry and wet deposition of SOx to the ambient air 

concentration of SOx, and TNOy equals the ratio of the combined dry and wet deposition of NOy

to the ambient air concentration of NOy.

As described in chapter 7 of the PA, reduced forms of nitrogen (NHx) are included in 

total nitrogen in the critical load equation, III-1. Reduced forms of nitrogen are treated 

separately, as are NOy and SOx, and the transference ratios are applied. This results in the 

following critical load expression that is defined explicitly in terms of the indicators NOy and 

SOx:

CLANClim(N + S) = ([BC]0

*

- [ANClim])Q + Neco = [NOy]TNOy + [SOx]TSOx + NHx (III-2)

This is the same equation as III-1, with the deposition associated with the critical load translated 

to deposition from ambient air concentrations via transference ratios. In addition, deposition of 

reduced nitrogen, oxidized nitrogen and oxidized sulfur are treated separately. 

Transference ratios are a modeled construct, and therefore cannot be compared directly to 

measurable quantities. There is an analogy to deposition velocity, as a transference ratio is 

basically an aggregated weighted average of the deposition velocities of all contributing species 
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across dry and wet deposition, and transference ratio units are expressed as distance/time. 

However, wet deposition commonly is not interpreted as the product of a concentration times a 

velocity. Direct wet deposition observations are available which integrate all of the processes, 

regardless of how well they may be understood, related to wet deposition into a measurable 

quantity. There are reasonable analogies between the processes governing dry and wet 

deposition, from a fundamental mass transfer perspective. In both cases there is a transfer of 

mass between the dry ambient phase and another medium, either a surface or vegetation in the 

case of dry deposition, or a rain droplet or cloud in the case of wet precipitation. The specific 

thermodynamic properties and chemical/biological reactions that govern the transfer of dry mass 

to plants or aqueous droplets differ, but either process can be based on conceptualizing the 

product of a concentration, or concentration difference, times a mass transfer coefficient which is 

analogous to the basic dry deposition model: dry deposition = concentration x velocity (US 

EPA, 2011, Appendix F). 

Transference ratios require estimates of wet deposition of NOy and SOx, dry deposition of 

NOy and SOx, and ambient air concentrations of NOy and SOx. Possible sources of information 

include model estimates or a combination of model estimates and observations, recognizing that 

dry deposition is a modeled quantity that can use observed or modeled estimates of 

concentration. The limited amount of NOy measurements in acid-sensitive areas as well as the 

combination of representative NOy, SO2 and SO4 observations generally preclude the use of 

observations for development of a standard that is applicable nationally.

The PA considers a blending of observations and models to take advantage of their 

relative strengths; e.g., combining the NADP wet deposition observations, modeled dry

deposition, and a mix of modeled and observed concentrations, using the model for those species 
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not measured or measured with very sparse spatial coverage. A potential disadvantage of mixing 

and matching observations and model estimates is to lose consistency afforded by using just 

modeling alone. A modeling platform like CMAQ is based on adhering to consistent treatment 

of mass conservation, by linking emission inputs with air concentrations and concentrations to 

deposition. Inconsistencies from combining processes from different analytical platforms 

increase the chance that mass (of nitrogen or sulfur) would unintentionally be increased or 

decreased as the internal checking that assures mass conservation is lost. Transference ratios 

incorporate a broad suite of atmospheric processes and consequently an analytical approach that 

instills consistency in the linkage of these processes is preferable to an approach lacking such 

inherent consistency. This contention does not mean that observations alone, if available, could 

not be used, but suggests that the inconsistencies in combining models and observations for the 

purposes of developing transference ratios has the potential for creating unintended artifacts. 

While there is a reasonable conceptual basis for the concept of an aggregated deposition 

velocity referred to in the PA as a transference ratio, there is very limited ability to compare 

observed and calculated ratios. This is because the deposition velocity is dependent on 

individual species, and the mass transfer processes of wet and dry removal, while conceptually

similar, are different. Consequently, there does not exist a meaningful approach to measure such 

an aggregated or lumped parameter. Therefore, at this time, the evaluation of transference ratios 

is based on sensitivity studies, analysis of variability, and comparisons with other models, as 

described in Appendix F of the PA. 

As discussed in Appendix F, the interannual variability, as well as the sensitivity to 

emission changes of roughly 50 percent, results in changes of transference ratios of 

approximately 5 to 10 percent. Part of the reason for this inherent stability is due to the co-
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dependence of concentration and deposition. For example, as concentrations are reduced as a 

result of emissions reductions, deposition in turn is reduced since deposition is a direct linear 

function of concentration leading to negligible impact on the deposition-to-concentration ratio. 

Likewise, an overestimate of concentration likely does not induce a bias in the transference ratio. 

While it is important to continue to improve the model’s ability to match ambient concentrations 

in time and space, the bias of a modeled estimate of concentration relative to observations does 

not necessarily result in a bias in a calculated transference ratio. In effect, this consideration of 

bias cancellation reduces the sensitivity of transference ratios to model uncertainties and affords 

increased confidence in the stability of these ratios. Based on the series of sensitivity and 

variability analyses, the PA concludes that the transference ratios are relatively stable and 

provide a sound metric for linking deposition and concentration.

As discussed in the PA, transference ratios are dependent on the platform upon which 

they are constructed. Comparisons of transference ratios constructed from different modeling 

platforms do exhibit significant differences. While this divergence of results may be explained 

by a variety of differences in process treatments, input fields and incommensurabilities in species 

definitions and spatial configurations, it does suggest two very important conclusions. First, the 

idea of using multiple platforms for different parts of the country may be problematic as there 

does not exist a reliable approach to judge acceptance which is almost always based on 

comparisons to observations. Second, since transference ratios are based on concentrations and 

deposition, as the uncertainties in each of those components are reduced, the relative uncertainty 

in the ratios also is reduced. This means that basic improvements in the model’s ability to 

reproduce observed wet deposition and ambient concentration fields enhance the relative 

confidence in the constructed transference ratios. Similarly, as in-situ dry deposition flux
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measurements become available that enable a more rigorous evaluation and diagnosis of 

modeled dry deposition processes, the expected improved treatment of dry deposition also would 

increase confidence in transference ratios. Finally, deposition is directly related to ambient air 

concentrations. Models like CMAQ rely on the concentration-to-deposition linkage to calculate 

deposition, which is the foundation for broadly based and robust assessments addressing 

atmospheric deposition. In principle, the use of a modeled constructed transference ratio is based 

on the same premise by which we use models to estimate deposition in the first place.

The shortage of widely available ambient air observations and the fact that estimates of 

dry deposition requires modeling, collectively suggests that a unified modeling platform is the 

best approach for constructing transference ratios. The PA (US EPA, 2011, section 2) considers 

CMAQ and other models, such as CAMx and Canada’s AURAMS - A Unified Regional Airquality Modeling System (Smythe et al., 2008), and concludes that CMAQ is the preferred 

modeling platform for constructing transference ratios. This conclusion reflects the view that for 

the purposes of defining transference ratios, a modeling platform should: (1) be a multiple 

pollutant model recognizing the myriad of connections across pollutant categories that directly

and indirectly impact nitrogen and sulfur characterization, (2) include the most comprehensive 

scientific treatments of atmospheric processes that relate directly and indirectly to characterizing 

concentrations and deposition, (3) have an infrastructure capability that accommodates the 

inclusion of improved scientific treatments of relevant processes and important input fields, and 

(4) undergo frequent reviews regarding the adequacy of the underlying science as well as the 

appropriateness in applications. The CMAQ platform exhibits all these characteristics. It has 

been (and continues to be) extensively evaluated for several pollutant categories, and is 

supported by a central infrastructure of EPA scientists, whose mission is to improve and evaluate 
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the CMAQ platform. More directly, CMAQ, and its predecessor versions, has a long track 

record going back to the NAPAP in the 1980’s of specific improvements in deposition processes, 

which are described in Appendix F of the PA.

4. Aquatic Acidification Index

Having established the various expressions that link atmospheric deposition of nitrogen 

and sulfur to ANC and the transference ratios that translate atmospheric concentrations to 

deposition of nitrogen and sulfur, the PA derived the following expression of these linkages, 

which separates reduced forms of nitrogen, NHx, from oxidized forms:

ANCcalc = {ANClim + CLr/Qr } – NHx/Qr - TNOy [NOy]/Qr - TSOx[SOx]/Qr (III-3)

Equation III-3 is the basic expression of the form of a standard that translates the 

conceptual framework into an explicit expression that defines ANC as a function of the ambient 

air indicators, NOy and SOx reduced nitrogen deposition9, and the critical load necessary to 

achieve a target ANC level. This equation calculates an expected ANC value based on ambient 

concentrations of NOy and SOx. The calculated ANC will differ from the target ANC (ANClim) 

depending on how much the nitrogen and sulfur deposition associated with NOy, SOx, and NHx

differs from the critical load associated with just achieving the target ANC. 

Based on equation III-3, the PA defines an AAI that is more simply stated using terms that 

highlight the ambient air indicators:

 

9 Because NHx is characterized directly as deposition, not as an ambient concentration in this equation, no 

transference ratio is needed for this term.
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AAI = F1 – F2 – F3[NOy] – F4[SOx] (III- 4)

where the AAI represents the long term (or steady state) ANC level associated with ambient air 

concentrations of NOy and SOz. The factors F1 through F4 convey three attributes: a relative 

measure of the ecosystem’s ability to neutralize acids (F1), the acidifying potential of reduced 

nitrogen deposition (F2), and the deposition-to-concentration translators for NOy (F3) and SOx 

(F4). 

Specifically:

F1 = ANClim + CLr/Qr ;

F2 = NHx/Qr = NHx deposition divided by Qr;

F3 = TNOy/ Qr ; TNOy is the transference ratio that converts ambient air concentrations of NOy to 

deposition of NOy; and

F4 = TSOx/ Qr ; TSOx is the transference ratio that converts ambient air concentrations of SOx to 

deposition of SOx.

All of these factors include representative Qr to maintain unit (and mass) consistency between 

the AAI and the terms on the right side of equation III-4.

The F1 factor is the target ANC level plus the amount of deposition (critical load) the 

ecosystem can receive and still achieve the target level. It incorporates an ecosystem’s ability to 

generate acid neutralizing capacity through base cation supply ([BC]*0) and to neutralize 

acidifying nitrogen deposition through Neco, both of which are incorporated in the CL term. As 

noted above, because Neco can only neutralize nitrogen deposition (oxidized or reduced) there 

may be rare cases where Neco exceeds the combination of reduced and oxidized nitrogen 

deposition. Consequently, to ensure that the AAI equation is applicable in all cases that may
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occur, equation III-4 is conditional on total nitrogen deposition, {NHx + F3[NOy]}, being 

greater than Neco. In rare cases where Neco is greater than {NHx + F3[NOy]}, F2, F3, and 

Neco would be set equal to 0 in the AAI equation. The consequence of setting F2 and F3 to zero 

is simply to constrain the AAI calculation just to SOx, as nitrogen would have no bearing on 

acidifying contributions in this case.

The PA concludes that equation III-4 (US EPA, 2011,equation 7-12), which defines the 

AAI, is ecologically relevant and appropriate for use as the form of a national standard designed 

to provide protection for aquatic ecosystems from the effects of acidifying deposition associated 

with concentrations of oxides of nitrogen and sulfur in the ambient air. This AAI equation does 

not, however, in itself, define the spatial areas over which the terms of the equation would apply. 

To specify values for factors F1 through F4, it is necessary to define spatial areas over which 

these factors are determined. Thus, it is necessary to identify an approach for spatially

aggregating water bodies into ecologically meaningful regions across the U.S., as discussed 

below.

5. Spatial Aggregation

As discussed in the PA, one of the unique aspects of this form is the need to consider the 

spatial areas over which values for the F factors in the AAI equation are quantified. Ecosystems 

across the U.S. exhibit a wide range of geological, hydrological and vegetation characteristics 

that influence greatly the ecosystem parameters, Q, BC0

*

and Neco that are incorporated in the 

AAI. Variations in ecosystem attributes naturally lead to wide variability in the sensitivities of 

water bodies in the U.S. to acidification, as well as in the responsiveness of water bodies to 

changes in acidifying deposition. Consequently, variations in ecosystem sensitivity, and the 

uncertainties inherent in characterizing these variations, must be taken into account in 
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developing a national standard. In developing a secondary NAAQS to protect public welfare, the 

focus of the PA is on protecting sensitive populations of water bodies, not on each individual 

water body, which is consistent with the Agency’s approach to protecting public health through 

primary NAAQS that focus on susceptible populations, not on each individual.

The approach used for defining ecologically relevant regions across the U.S. in the PA

(US EPA, 2011, section 7.2.5) is described below, along with approaches to characterizing each 

region as acid sensitive or relatively non-acid sensitive. This characterization facilitates a more 

detailed analysis and focus on those regions that are relatively more acid sensitive. This 

characterization is also used to avoid over-protection in relatively non-acid sensitive regions, 

regions that would receive limited benefit from reductions in the deposition of oxides of nitrogen 

and sulfur with respect to aquatic acidification effects. Approaches to developing representative 

values for each of the terms in the AAI equation (factors F1 through F4) for each ecologically

relevant region for which sufficient data are available are then discussed. These spatial

aggregation approaches are generally applicable to the contiguous U.S.. The following 

discussion also addresses the development of factors for data-limited regions and specifically for 

Hawaii, Alaska and the U.S. territories.

Stated more simply, this section discusses appropriate ways to divide the country into 

ecologically relevant regions; to characterize each region as either acid sensitive or relatively 

non-acid sensitive; and to determine values of factors F1 through F4 for each region, taking into 

consideration the acid sensitivity of each region and the availability of relevant data. For each 

such region, the AAI would be calculated based on the values of factors F1 through F4 specified 

for that region.
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In considering approaches to spatial aggregation, the PA focuses on methods that have 

been developed to define ecologically relevant regions, referred to as ecoregions, which are 

meaningfully related to the factors that are relevant to aquatic acidification. As noted above, the 

PA did not focus on looking at each individual water body, nor did it focus on aggregating over 

the entire nation, which would preclude taking into account the inherent variability in 

atmospheric and ecological factors that fundamentally modify the relationships that are central to 

the development of an ecologically relevant AAI. 

Based on considering available classification schemes, the PA concludes that Omernik’s 

ecoregion classification (as described at http://www.epa.gov/wed/pages/ecoregions) is the most 

appropriate method to consider for the purposes of this review. This classification offers several 

levels of spatial delineation, has undergone an extensive scientific peer review process, and has 

explicitly been applied to delineating acid sensitive areas within the U.S. Further, the PA

concludes that ecoregion level III (Figure III-1) resolution, with 84 defined ecoregions in the 

contiguous U.S.,10 is the most appropriate level to consider for this purpose. The spatial 

resolution afforded by level III strikes an appropriate balance relative to the reasoning that 

supports conclusions on indicators, as discussed above. The PA concludes that the most detailed 

level of resolution (level IV) is not appropriate given the limited data availability to address 

nearly 1000 subdivisons within that level and the currently evolving nature of level IV regions. 

Further, level III ecoregions are preferred to level II in that level III ecoregions, but not level II 

ecoregions, are largely contiguous in space which allows for a more coherent development of 

information to quantify the AAI factors and to characterize the concentrations of NOy and SOx 

in the ambient air within each ecoregion.

 

10 We note that an 85th area within Omernik’s Ecoregion Level III is currently being developed for California.
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Appendix C of the PA includes a description of each level III ecoregion. The PA notes 

that the use of ecoregions is an appropriate spatial aggregation scheme for an AAI-based

standard focused on deposition-related aquatic acidification effects, while many of the same 

ecoregion attributes may be applicable in subsequent NAAQS reviews that may address other 

deposition-related aquatic and terrestrial ecological effects. Because atmospheric deposition is 

modified by ecosystem attributes, the types of vegetation, soils, bedrock geology, and 

topographic features that are the basis of this ecoregion classification approach also will likely be 

key attributes for other deposition-related effects (e.g., terrestrial acidification, nutrient 

enrichment) that link atmospheric concentrations to an aquatic or terrestrial ecological indicator.

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

The PA used Omernik’s original alkalinity data (US EPA, 2011, section 2) and more 

recent ANC data to delineate two broad groupings of ecoregions: acid-sensitive and relatively 

non-acid sensitive ecoregions. This delineation was made to facilitate greater focus on those 

ecoregions with water bodies that generally have greater acid sensitivity and to avoid overprotection in regions with generally less sensitive water bodies. The approach used to delineate 

acid-sensitive and relatively non-acid sensitive regions included an initial numerical-based 

sorting scheme using ANC data, which categorized ecoregions with relatively high ANC values 

as being relatively non-acid sensitive. This initial delineation resulted in 29 of the 84 Omernik 

ecoregions being categorized as acid sensitive. Subsequently, land use data were also considered 

to determine to what extent an ecoregion is of a relatively pristine and rural nature by quantifying 

the degree to which active management practices related to development and agriculture occur in 

each ecoregion. 

The overall objective is to produce a logical and practical grouping of ecoregions that 

experience adverse conditions with respect to aquatic acidification and are likely to respond to 

changes in concentrations of NOy and SOx in the ambient air and to the related deposition levels. 

To achieve this goal, a two-step process has been applied, first identifying acid sensitive 

ecoregions based on water quality data alone, and second identifying among those acid-sensitive 

ecoregions those with highly developed and managed areas. These highly developed and 

managed ecoregions are placed in a non-acid sensitive category to avoid over protection beyond 

what is requisite to protect public welfare. More specifically, in determining an ecoregion’s acid 

sensitivity status in step 1, ANC data across the 84 ecoregions are sorted (US EPA, 2011, section
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7 ) to determine the number of water bodies within a region with ANC values suggestive of acid 

sensitivity, so as to screen out regions with an overabundance of high ANC values. In reviewing 

the ANC data, the PA identified 29 ecoregions that meet two criteria: (1) greater than 5 percent 

of water bodies with data with ANC values less than 200 μeq/L and (2) greater than 1 percent of 

water bodies with ANC values less than 100 μeq/L. In step 2, land use data were used to identify

those acid sensitive ecoregions with significant managed areas that would not be considered as 

having a relatively pristine and rural character. The percentage of the combination of developed 

(residential, transportation, industrial and commercial) and agricultural (croplands, pastures, 

orchards, vineyards) land use was used as an indicator of managed land use area. Forest cover 

was used as an indicator of non-managed land use more directly reflecting the pristine quality of 

a region. Based on the 2006 National Land Cover Data base (NLCD, 

http://www.epa.gov/mrlc/nlcd-2006.html), acid sensitive ecoregions would meet both of the 

following land use data criteria: percent of developed and agricultural area less than 20 percent

combined with forested area greater than 50 percent. The combination of steps 1 and 2 identify

22 relatively acid sensitive areas (Table III-1 and Figure III-2). 

Consideration was also given to the use of naturally acidic conditions in defining 

relatively non-acid sensitive areas. For example, several of the ecoregions located in plains near 

the coast exhibit elevated dissolved organic carbon (DOC) levels, which is associated with 

naturally acidic conditions. The DOC in surface waters is derived from a variety of weak 

organic acid compounds generated from the natural availability and decomposition of organic 

matter from biota. Consequently, high DOC is associated with “natural” acidity, with the 

implication that a standard intended to protect against atmospheric contributions to acidity is not 

an area of focus. The evidence suggests that several of the more highly managed ecoregions in 
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coastal or near coastal transition zones are associated with relatively high DOC values, typically 

exceeding on average 5 mg/l, compared to other acid sensitive areas. Although there is sound 

logic to interpret naturally acidic areas as relatively non-acid sensitive, natural acidity indicators 

were not explicitly included in defining relatively non-acid sensitive areas as there does not exist 

a consensus-based quantifiable scientific definition of natural acidity. Approaches to explicitly

define natural acidity likely will be pursued in future reviews of the standard.

Table III-1. List of 22 Acid-Sensitive Areas

Ecoregion name Ecoregion number

Ridge and Valley 8.4.1

Northern Appalachian Plateau and Uplands 8.1.3

Piedmont 8.3.4

Western Allegheny Plateau 8.4.3

Southwestern Appalachians 8.4.9

Boston Mountains 8.4.6

Blue Ridge 8.4.4

Ouachita Mountains 8.4.8

Central Appalachians 8.4.2

Northern Lakes and Forests 5.2.1

Maine/New Brunswick Plains and Hills 8.1.8

North Central Appalachians 5.3.3

Northern Appalachian and Atlantic Maritime Highlands 5.3.1

Columbia Mountains/Northern Rockies 6.2.3

Middle Rockies 6.2.10

Wasatch and Uinta Mountains 6.2.13

North Cascades 6.2.5

Cascades 6.2.7

Southern Rockies 6.2.14

Sierra Nevada 6.2.12

Idaho Batholith 6.2.15

Canadian Rockies 6.2.4
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Figure III-2. Acid-sensitive ecoregions identified in grey fill (22 out of 84 ecoregions).

b. Representative Ecoregion-specific Factors

Having concluded that the Omernik level III ecoregions are an appropriate approach to 

spatial aggregation for the purpose of a standard to address deposition-related aquatic 

acidification effects, the PA uses those ecoregions to define each of the factors in the AAI 

equation. As discussed below, factors F1 through F4 in equation III-4 are defined for each 

ecoregion by specifying ecoregion-specific values for each factor based on monitored or 

modeled data that are representative of each ecoregion.
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i. Factor F1

As discussed above, factor F1 reflects a relative measure of an ecosystem’s ability to 

neutralize acidifying deposition, and is defined as: F1 = ANClim + CLr/Qr. The value of F1 for 

each ecoregion would be based on a representative critical load for the ecoregion (CLr) 

associated with a single national target ANC level (ANClim, discussed below in section III.D), 

as well as on a representative runoff rate (Qr). To specify ecoregion-specific values for the term 

Qr, the PA used the median value of the distribution of Q values that are available for water 

bodies within each ecoregion. To specify ecoregion-specific representative values for the term 

CLr in factor F1, a distribution11 of calculated critical loads was created for the water bodies in 

each ecoregion for which sufficient water quality and hydrology data are available.12 The 

representative critical load was then defined to be a specific percentile of the distribution of 

critical loads in the ecoregion. Thus, for example, using the 90th percentile means that within an 

ecoregion, 90 percent of the water bodies would be expected to have higher calculated critical 

loads than the representative critical load. That is, if the representative critical load were to 

occur across the ecoregion, 90 percent of the water bodies would be expected to achieve the 

national ANC target or better.

The specific percentile selected as part of the definition of F1 is an important parameter 

that directly impacts the representative critical load specified for each ecoregion, and therefore 

the degree of protectiveness of the standard. A higher percentile corresponds to a lower critical 

load and, therefore, to lower allowable ambient air concentrations of NOy and SOx and related 

 

11 The distribution of critical loads was based on CL values calculated with Neco at the lake level. Consideration 

could also be given to using a distribution of CLs without Neco and adding the ecoregion average Neco value to the 

nth percentile critical load. This would avoid cases where the lake-level Neco value potentially could be greater 

than total nitrogen deposition. The CL at the lake level represents the CL for the lake to achieve the specified 

national target ANC value.

12 The PA judged the data to be sufficient for this purpose if data are available from more than 10 water bodies in an 

ecoregion.
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deposition to achieve a target AAI level. In conjunction with the other terms in the AAI

equation, alternative forms can be appropriately characterized in part by identifying a range of 

alternative percentiles. The choice of an appropriate range of percentiles to consider for acidsensitive and relatively non-acid sensitive ecoregions, respectively, is discussed below. 

(a) Acid-sensitive Ecoregions

In identifying percentiles that are appropriate to consider for the purpose of calculating 

factor F1 for ecoregions characterized as acid sensitive, the PA concludes that it is appropriate to 

focus on the lower (more sensitive) part of the distribution of critical loads, so as to ensure that 

the ecoregion would be represented by relatively more acid sensitive water bodies within the 

ecoregion. Specifying factor F1 in this way would help to define a standard that would be 

protective of the population of acid sensitive water bodies within an ecoregion, recognizing that 

even ecoregions characterized as acid sensitive may contain a number of individual water bodies 

that are not acid sensitive. The PA recognizes that there is no basis for independently evaluating 

the degree of protectiveness afforded by any specific percentile value, since it is the combination 

of form and level, in conjunction with the indicator and averaging time, which determine the 

degree of protectiveness of a standard. In light of this, the PA concludes that it is appropriate to 

consider a range of percentiles, from well above the 50th percentile, or median, of the distribution 

to somewhat below the highest value (in terms of sensitivity; a high degree of sensitivity

corresponds to a low value for critical load). More specifically, the PA concludes it is 

appropriate to consider percentiles in the range of the 70th to the 90th percentile (of sensitivity). 

This conclusion is based on the judgment that it would not be appropriate to represent an 

ecoregion with the lowest or near lowest critical load, so as to avoid potential extreme outliers 

that can be seen to exist at the extreme end of the data distributions, which would not be 
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representative of the population of acid sensitive water bodies within the ecoregion and could 

lead to an overly protective standard. At the same time, in considering ecoregions that are 

inherently acid sensitive, it is judged to be appropriate to limit the lower end of the range for 

consideration to the 70th percentile, a value well above the median of the distribution, so that a 

substantial majority of acid-sensitive water bodies are protected. 

In considering this conclusion, the CASAC Panel noted that the data bases for calculating 

critical loads within an ecoregion are not necessarily representative of all water bodies within an 

ecoregion. That is, in many ecoregions the lake sampling design used in studies that generated 

the relevant data may have focused on the relatively more sensitive water bodies within an 

ecoregion (Russell and Samet, 2011a). Consequently, a given percentile of the distribution of 

calculated critical loads, based on sampled water bodies, may not be representative of that 

percentile of all water bodies across an entire ecoregion. To the extent that the sampling of water 

bodies within an ecoregion was skewed toward the relatively more sensitive water bodies, 

selecting a given percentile from the distribution of available critical loads would result in a 

somewhat higher percentile of all water bodies within that ecoregion having a higher calculated 

critical load than the representative critical load value. Thus, the extent to which study sampling 

designs have resulted in skewed distributions of calculated critical loads is an uncertainty that is 

appropriate to consider in selecting a percentile for the purpose of defining the factor F1 in the 

AAI equation. 

(b) Non-acid sensitive Ecoregions

With regard to identifying percentiles that are appropriate to consider for the purpose of 

calculating factor F1 for ecoregions characterized as relatively non-acid sensitive, the PA

recognizes that while such ecoregions are generally less sensitive to acidifying deposition from 
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oxides of nitrogen and sulfur, they may contain a number of water bodies that are acid sensitive 

This category includes ecoregions that are well protected from acidification effects due to natural 

production of base cations and high ANC levels, as well as naturally acidic systems with limited 

base cation production and consequently very low critical loads. Therefore, the use of a critical 

load that would be associated with a highly sensitive water body in a naturally acidic system 

would impose a high degree of relative protection in terms of allowable ambient air 

concentrations of oxides of nitrogen and sulfur and related deposition, while potentially

affording little or no public welfare benefit from attempting to improve a naturally acidic system.

Based on these considerations, the PA concludes it is appropriate to consider the use of a 

range of percentiles that extends lower than the range identified above for acid-sensitive 

ecoregions. Consideration of a lower percentile would avoid representing a relatively non-acid 

sensitive ecoregion by a critical load associated with relatively more acid-sensitive water bodies. 

In particular, the PA concludes it is appropriate to focus on the median or 50th percentile of the 

distribution of critical loads so as to avoid over-protection in such ecoregions. Recognizing that 

relatively non-acid sensitive ecoregions generally are not sampled to the extent that acidsensitive ecoregions are, it also is appropriate to consider using the median critical load of all 

relatively non-acid sensitive ecoregions for each such ecoregion.

ii. Factor F2

As discussed above, factor F2 is the amount of reduced nitrogen deposition within an 

ecoregion, including the deposition of both ammonia gas and ammonium ion, and is defined as: 

F2 = NHx/Qr. The PA calculated the representative runoff rate, Qr, using a similar approach as 

noted above for factor F1; i.e., the median value of the distribution of Q values that are available 

for water bodies within each ecoregion. In the PA, 2005 CMAQ model simulations over 12-km 
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grids are used to calculate an average value of NHx for each ecoregion. The NHx term is based 

on annual average model outputs for each grid cell, which are spatially averaged across all the 

grid cells contained in each ecoregion to calculate a representative annual average value for each 

ecoregion. The PA concludes that this approach of using spatially averaged values is appropriate 

for modeling, largely due to the relatively rapid mixing of air masses that typically results in 

relatively homogeneous air quality patterns for regionally dispersed pollutants. In addition, there 

is greater confidence in using spatially averaged modeled atmospheric fields than in using 

modeled point-specific fields.

This averaging approach is also used for the air concentration and deposition terms in 

factors F3 and F4, as discussed below. The PA notes that modeled NHx deposition exhibits 

greater spatial variability than the other modeled terms in factors F3 and F4. Recognizing this 

greater variability, the PA concludes that it would be appropriate to consider alternative 

approaches to specifying the value of NHx. One such approach might involve the use of more 

localized and/or contemporaneous modeling in areas where this term is likely to be particularly 

variable and important.

iii. Factors F3 and F4

As discussed above, factors F3 and F4 are the ratios that relate ambient air concentrations 

of NOy and SOx to the associated deposition, and are defined as follow: F3 = TNOy/ Qr and F4 

= TSOx/ Qr. TNOy is the transference ratio that converts ambient air concentrations of NOy to 

deposition of NOy and TSOx is the transference ratio that converts ambient air concentrations of 

SOx to deposition of SOx. The representative runoff rate, Qr, is calculated as for factors F1 and 

F2. The transference ratios are based on the 2005 CMAQ simulations, using average values for 

each ecoregion, as noted above for factor F2. More specifically, the transference ratios are 
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calculated as the annual deposition of NOy or SOx spatially averaged across the ecoregion and 

divided by the annual ambient air concentration of NOy or SOx, respectively, spatially averaged 

across the ecoregion.

c. Factors in Data-limited Ecoregions 

As discussed above in section III.B.5.a, in the PA the initial delineation of acid-sensitive 

and relatively non-acid sensitive ecoregions was based on available ANC and alkalinity data. 

Areas not meeting the ANC criteria described above are categorized as relatively non-acid 

sensitive. The development of a reasonable distribution of critical loads for water bodies within 

an ecoregion for the purpose of identifying the representative critical load requires additional 

data, including more specific water quality data for major cations and anions. This means that 

the water bodies that can be used to develop a distribution of critical loads is generally a subset 

of those water bodies for which ANC data are available Consequently, there are certain 

ecoregions with sparse data that are not suitable for developing a distribution of critical loads.

As noted above, the PA judges that it is not appropriate to develop such distributions 

based on data from less than ten water bodies within an ecoregion. Twelve such ecoregions, 

which included only relatively non-acid sensitive ecoregions, were characterized as being datalimited. For these ecoregions, the PA considered alternative approaches to specifying values for 

the terms CLr and Qr for the purpose of determining values for each of the factors in the AAI 

equation. For these data-limited ecoregions, the PA judges that it is appropriate to use the 

median values of CLr and Qr from the distributions of these terms for all other relatively non-acid 

sensitive ecoregions, rather than attempting to use severely limited data to develop a value for 

these terms based solely on data from such an ecoregion. Further, consideration could be given 

to using a single national default value for all relatively non-acid sensitive ecoregions. The PA
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notes that this data limitation is not a concern in specifying values for the other terms in the AAI 

equation for such ecoregions, since those terms are based on data from the 2005 CMAQ model 

simulation, which covers all ecoregions across the contiguous U.S.

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

The above methods for specifying ecoregion-specific values for the factors in the AAI 

equation apply to those ecoregions within the contiguous U.S. For areas outside the continental 

U.S., including Hawaii, Alaska, and the U.S. Territories, there is currently a lack of available 

data to characterize the sensitivity of such areas, as well as a lack of water body-specific data and 

CMAQ-type modeling to specify values for the F1 through F4 factors. Thus, the PA has 

considered possible alternative approaches to specifying values for factors F1 through F4 in the 

AAI equation for these areas.

One such approach could be to specify area-specific values for the factors based on 

values derived for ecoregions with similar acid sensitivities, to the extent that relevant 

information can be obtained to determine such similarities. Such an approach would involve 

conducting an analysis to characterize similarities in relevant ecological attributes between 

ecoregions in the contiguous U.S. and these areas outside the contiguous U.S. so as to determine 

the appropriateness of utilizing ecoregion-specific values for the CLr and Qr terms from one or 

more ecoregions within the contiguous U.S. This approach would also involve conducting 

additional air quality modeling for the areas that are outside the geographical scope of the 

currently available CMAQ model simulations, so as to develop the other information necessary 

to specify values for factors F2 through F4 for these areas.

A second approach could rely on future data collection efforts to establish relevant 

ecological data within these areas that, together with additional air quality modeling, could be 
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used to specify area-specific values for factors F1 through F4. Until such time as relevant data 

become available, these areas could be treated the same as data-limited ecoregions in the 

contiguous U.S. that are relatively non-acid sensitive.

The PA concludes that either approach would introduce substantial uncertainties that 

arise from attempting to extrapolate values based on similarity assumptions or arbitrarily 

assigning values for factors in the AAI equation that would be applicable to these areas outside 

the contiguous U.S. In light of such uncertainties, the PA concludes that it would also be 

appropriate to consider relying on the existing NO2 and SO2 secondary standards in these areas 

for protection of any potential direct or deposition-related ecological effects that may be 

associated with the presence of oxides of nitrogen and sulfur in the ambient air. The PA

concludes that relying on existing secondary standards in these areas is preferable to using a 

highly uncertain approach to allow for the application of a new standard based on the AAI in the 

absence of relevant area-specific data.

6. Summary of the AAI Form

With regard to the form of a multi-pollutant air quality standard to address depositionrelated aquatic acidification effects, the PA concludes that consideration should be given to an 

ecologically relevant form that characterizes the relationships between the ambient air indicators 

for oxides of nitrogen and sulfur, the related deposition of nitrogen and sulfur, and the associated 

aquatic acidification effects in terms of a relevant ecological indicator. Based on the available 

information and assessments, consideration should be given to using ANC as the most 

appropriate ecological indicator for this purpose, in that it provides the most stable metric that is 

highly associated with the water quality properties that are directly responsible for the principal 
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adverse effects associated with aquatic acidification: fish mortality and reduced aquatic species 

diversity.

The PA developed such a form, using a simple equation to calculate an AAI value in 

terms of the ambient air indicators of oxides and nitrogen and sulfur and the relevant ecological 

and atmospheric factors that modify the relationships between the ambient air indicators and 

ANC. Recognizing the spatial variability of such factors across the U.S., the PA concludes it is 

appropriate to divide the country into ecologically relevant regions, characterized as acidsensitive or relatively non-acid-sensitive, and specify the value of each of the factors in the AAI 

equation for each such region. Omernik ecoregions, level III, are identified as the appropriate set 

of regions over which to define the AAI. There are 84 such ecoregions that cover the continental 

U.S. This set of ecoregions is based on grouping a variety of vegetation, geological, and 

hydrological attributes that are directly relevant to aquatic acidification assessments and that 

allow for a practical application of an aquatic acidification standard on a national scale.

The PA defines AAI by the following equation: AAI = F1 – F2 – F3[NOy] – F4[SOx]. 

Factors F1 through F4 would be defined for each ecoregion by specifying ecoregion-specific 

values for each factor based on monitored or modeled data that are representative of each 

ecoregion. The F1 factor is also defined by a target ANC value. More specifically:

(1) F1 reflects a relative measure of an ecosystem’s ability to neutralize acidifying 

deposition. The value of F1 for each ecoregion would be based on a representative 

critical load for the ecoregion associated with a single national target ANC level, as well 

as on a representative runoff rate. The representative runoff rate, which is also used in 

specifying values for the other factors, would be the median value of the distributions of 

runoff rates within the ecoregion. The representative critical load would be derived from 
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a distribution of critical loads calculated for each water body in the ecoregion for which 

sufficient water quality and hydrology data are available. The representative critical load 

would be defined by selecting a specific percentile of the distribution.

In identifying a range of percentiles that are appropriate to consider for this purpose, 

regions categorized as acid sensitive were considered separately from regions categorized 

as relatively non-acid sensitive. For acid sensitive regions, the PA concludes that 

consideration should be given to selecting a percentile from within the range of the 70th to 

the 90th percentile. The lower end of this range was selected to be appreciably above the 

median value so as to ensure that the critical load would be representative of the 

population of relatively more acid sensitive water bodies within the region, while the 

upper end was selected to avoid the use of a critical load from the extreme tail of the 

distribution which is subject to a high degree of variability and potential outliers. For 

relatively non-acid sensitive regions, the PA concludes that consideration should be given 

to selecting the 50th percentile to best represent the distribution of water bodies within 

such a region, or alternatively to using the median critical load of all relatively non-acid 

sensitive areas, recognizing that such areas are far less frequently evaluated than acid 

sensitive areas. Using either of these approaches would avoid characterizing a generally

non-acid-sensitive region with a critical load that is representative of relatively acid 

sensitive water bodies that may exist within a generally non-acid sensitive region.

(2) F2 reflects the deposition of reduced nitrogen. Consideration should be given to 

specifying the value of F2 for each region based on the averaged modeled value across 

the region, using national CMAQ modeling that has been conducted by EPA. 

Consideration could also be given to alternative approaches to specifying this value, such 
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as the use of more localized and/or contemporaneous modeling in areas where this term is 

likely to be particularly variable and important. 

(3) F3 and F4 reflect transference ratios that convert ambient air concentrations of NOy

and SOx, respectively, into related deposition of nitrogen and sulfur. Consideration 

should be given to specifying the values for F3 and F4 for each region based on CMAQ 

modeling results averaged across the region. We conclude that specifying the values or 

the transference ratios based on CMAQ modeling results alone is preferred to an 

alternative approach that combines CMAQ model estimates with observational data.

(4) The terms [NOy] and [SOx] reflect ambient air concentrations measured at 

monitoring sites within each region.

Using the equation, a value of AAI can be calculated for any measured values of ambient 

NOy and SOx. For such a NAAQS, the Administrator would set a single, national value for the 

level of the AAI used to determine achievement of the NAAQS, as discussed below in section 

III.D. The ecoregion-specific values for factors F1 through F4 would be specified by EPA based 

on the most recent data and CMAQ model simulations, and codified as part of such a standard. 

These factors would be reviewed and updated as appropriate in the context of each periodic 

review of the NAAQS. 

The PA developed specific F factors for each ecoregion based on the approach discussed 

above, using alternative percentiles and alternative national target ANC levels. The results of 

this analysis for ecoregions characterized as acid sensitive are presented in Table 7-1a-d in the 

PA. 

C. Averaging Time
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As discussed in section 7.3 of the PA, aquatic acidification can occur over both long- and 

short-term timescales. Long-term cumulative deposition of nitrogen and sulfur is reflected in the 

chronic acid-base balance of surface waters as indicated by measured annual ANC levels. 

Similarly, the use of steady state critical load modeling, which generates critical loads in terms of 

annual cumulative deposition of nitrogen and sulfur, means that the focus of ecological effects 

studies based on critical loads is on the long-term equilibrium status of water quality in aquatic 

ecosystems. Much of the evidence of adverse ecological effects associated with aquatic 

acidification, as discussed above in section II.A, is associated with chronically low ANC levels. 

Protection against a chronic ANC level that is too low is provided by reducing overall annual 

average deposition levels for nitrogen and sulfur.

Reflecting this focus on long-term acidifying deposition, the PA developed the AAI that 

links ambient air indicators to deposition-related ecological effects, in terms of several factors, 

F1 through F4. As discussed above, these factors are all calculated as annual average values, 

whether based on water quality and hydrology data or on CMAQ model simulations. In the 

context of a standard defined in terms of the AAI, the PA concludes that it is appropriate to 

consider the same annual averaging time for the ambient air indicators as is used for the factors 

in the AAI equation.

We also recognize that short-term (i.e., hours or days) episodic changes in water 

chemistry, often due to changes in the hydrologic flow paths, can have important biological 

effects in aquatic ecosystems. Such short-term changes in water chemistry are termed “episodic 

acidification.” Some streams may have chronic or base flow chemistry that is generally healthy 

for aquatic biota, but may be subject to occasional acidic episodes with potentially lethal 
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consequences. Thus, short-term episodic ecological effects can occur even in the absence of 

long-term chronic acidification effects.

Episodic declines in pH and ANC are nearly ubiquitous in drainage waters throughout the 

eastern U.S. Episodic acidification can result from several mechanisms related to changes in 

hydrologic flow paths. For example, snow can store nitrogen deposited throughout the winter 

and snowmelt can then release this stored nitrogen, together with nitrogen derived from 

nitrification in the soil itself, in a pulse that leads to episodic acidification in the absence of 

increased deposition during the actual episodic acidification event. The PA notes that inputs of 

nitrogen and sulfur from snowpack and atmospheric deposition largely cycle through soil. As a 

result, short-term direct deposition inputs are not necessarily important in episodic acidification. 

Thus, as noted in chapter 3 of the ISA, protection against episodic acidity events can be achieved 

by establishing a higher chronic ANC level.

Taken together, the above considerations support the conclusion that it is appropriate to 

consider the use of a long-term average for the ambient air indicators NOy and SOx for an aquatic 

acidification standard defined in terms of the AAI. The use of an annual averaging time for NOy

and SOx concentrations would be appropriate to provide protection against low chronic ANC 

levels, which in turn would protect against both long-term acidification and acute acidic 

episodes.

The PA has also considered interannual variability in both ambient air quality and in 

precipitation, which is directly related to the deposition of oxides of nitrogen and sulfur from the 

ambient air. While ambient air concentrations show year-to-year variability, often the year-toyear variability in precipitation is considerably greater, given the highly stochastic nature of 

precipitation. The use of multiple years over which annual averages are determined would 
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dampen the effects of interannual variability in both air quality and precipitation. For the 

ambient air indicators, the use of multiple-year averages would also add stability to calculations 

used to judge whether an area meets a standard defined in terms of the AAI. Consequently, the 

PA concludes that an annual averaging time based on the average of each year over a 

consecutive 3- to 5-year period is appropriate to consider for the ambient air indicators NOy and 

SOx. In reaching this conclusion, the PA notes that in its comments on the second draft PA, 

CASAC agreed that a 3- to 5-year averaging time was appropriate to consider (Russell and 

Samet, 2010b).

D. Level

As discussed above, the PA concludes that ANC is the ecological indicator best suited to 

reflect the sensitivity of aquatic ecosystems to acidifying deposition from oxides of nitrogen and 

sulfur in the ambient air. The ANC is an indicator of the aquatic acidification expected to occur 

given the natural buffering capacity of an ecosystem and the loadings of nitrogen and sulfur 

resulting from atmospheric deposition. Thus, the PA developed a new standard for aquatic 

acidification that is based on the use of chronic ANC as the ecological indicator as a component 

in the AAI.

The level of the standard would be defined in terms of a single, national value of the 

AAI. The standard would be met at a monitoring site when the multi-year average of the 

calculated annual values of the AAI was equal to or above the specified level of the standard.13 

The annual values of the AAI would be calculated based on the AAI equation using the assigned 

ecoregion-specific values for factors F1 through F4 and monitored annual average NOy and SOx

concentrations. Since the AAI equation is based on chronic ANC as the ecological indicator, the 

 

13 Unlike other NAAQS, where the standard is met when the relevant value is at or below the level of the standard 

since a lower standard level is more protective, in this case a higher standard level is more protective.
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level chosen for the standard would reflect a target chronic ANC value. As noted above, the 

assigned F factors for each ecoregion would be determined by EPA in the rulemaking to set the 

NAAQS, based on water quality and hydrology data, CMAQ modeling, the selected percentile 

that is used to identify a representative critical load within the ecoregion, and the selected level 

of the standard. The combination of the form of the standard, discussed above in section III.B, 

defined by the AAI equation and the assigned values of the F factors in the equation, other 

elements of the standard including the ambient air indicators (section III.A) and their averaging 

time (section III.C), and the level of the standard determines the allowable levels of NOy and SOx

in the ambient air within each ecoregion. All of the elements of the standard together determine 

the degree of protection from adverse aquatic acidification effects associated with oxides of 

nitrogen and sulfur in the ambient air. The level of the standard plays a central role in 

determining the degree of protection provided and is discussed below.

The PA focuses primarily on information that relates degrees of biological impairment 

associated with adverse ecological effects to aquatic ecosystems to alternative levels of ANC in 

reaching conclusions regarding the range of target ANC levels that is appropriate to consider for 

the level of the standard. The PA develops the rationale for identifying a range of target ANC 

levels that is appropriate to consider by addressing questions related to the following areas: (1) 

associations between ANC and pH levels to provide an initial bounding for the range of ANC 

values to be considered; (2) evidence that allows for the delineation of specific ANC ranges 

associated with varying degrees of severity of biological impairment ecological effects; (3) the 

role of ANC in affording protection against episodic acidity; (4) implications of the time lag 

response of ANC to changes in deposition; (5) past and current examples of target ANC values 
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applied in environmental management practices; and (6) data linking public welfare benefits and 

ANC.

1. Association Between pH Levels and Target ANC Levels 

As discussed above in section II.A and more fully in chapter 3 of the PA, specific levels 

of ANC are associated with differing levels of risk of biological impairment in aquatic 

ecosystems, with higher levels of ANC resulting in lower risk of ecosystem impacts, and lower 

ANC levels resulting in risk of both higher intensity of impacts and a broader set of impacts. 

While ANC is not the causal agent determining biological effects in aquatic ecosystem, it is a 

useful metric for determining the level at which a water body is protected against risks of 

acidification. There is a direct correlation between ANC and pH levels which, along with 

dissolved aluminum, are more closely linked to the biological causes of ecosystem response to 

acidification. 

Because there is a direct correlation between ANC and pH levels, the selection of target 

ANC levels is informed in part through information on effects of pH as well as direct studies of 

effects related to ANC. Levels of pH are closely associated with ANC in the pH range of 

approximately 4.5 to 7. Within this range, higher ANC levels are associated with higher pH 

levels. At a pH level of approximately 4.5, further reductions in ANC generally do not correlate 

with pH, as pH levels remain at approximately 4.5 while ANC values fall substantially. 

Similarly, at a pH value of approximately 7, ANC values can continue to increase with no 

corresponding increase in pH. As pH is the primary causal indicator of effects related to aquatic 

acidification, this suggests that ANC values below approximately -50 μeq/L (the apparent point 

in the relationship between pH and ANC where pH reaches a minimum) are not likely to result in 

further damage. In addition, ANC values around and above approximately 100 μeq/L (the 
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apparent region in the relationship where pH reaches a maximum) are not likely to confer 

additional protection. As a result, the initial focus in the PA was on target ANC values in the 

range of -50 to 100 μeq/L.

2. ANC Levels Related to Effects on Aquatic Ecosystems 

As discussed above in section II.A, the number of fish species present in a water body has 

been shown to be positively correlated with the ANC level in the water, with higher values 

supporting a greater richness and diversity of fish species. The diversity and distribution of 

phyto-zooplankton communities also are positively correlated with ANC. 

A summary of effects related to ANC ranges is shown above in Table II-1. Within the 

ANC range from approximately -50 to 100 μeq/L, linear and sigmoidal relationships are 

observed between ANC and ecosystem effects. On average, fish species richness is lower by one 

fish species for every 21 μeq/L decrease in ANC in Shenandoah National Park streams (ISA, 

section 3.2.3.4). As shown in Table II-1, ANC levels have been grouped into five categories 

related to expected ecological effects, including categories of acute concern ( <0 μeq/L), severe 

concern (0-20 μeq/L), elevated concern (20-50 μeq/L), moderate concern (50-100 μeq/L), and 

low concern (>100 μeq/L). This categorization is supported by a large body of research 

completed throughout the eastern U.S. (Sullivan et al., 2006).

Water bodies with ANC values less than or equal to 0 μeq/L at based flow are chronically 

acidic. Such ANC levels can lead to complete loss of species and major changes in the ability of 

water bodies to support diverse biota, especially in water bodies that are highly sensitive to 

episodic acidification. Based on the above considerations, the PA has focused on target ANC 

levels no lower than 0 μeq/L.
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As discussed in the PA, biota generally are not harmed when ANC values are >100 

μeq/L, due to the low probability that pH levels will be below 7. In the Adirondacks, the number 

of fish species also peaks at ANC values >100 μeq/L. This suggests that at ANC levels greater 

than 100 μeq/L, little risk from acidification exists in many aquatic ecosystems. At ANC levels 

below 100 μeq/L, overall health of aquatic communities can be maintained, although fish fitness 

and community diversity begin to decline. At ANC levels ranging from 100 down to 50 μeq/L, 

there is increasing likelihood that the fitness of sensitive species (e.g., brook trout, zooplankton) 

will begin to decline. When ANC concentrations are below 50 μeq/L, the probability of 

acidification increases substantially, and negative effects on aquatic biota are observed, including 

large reductions in diversity of fish species and changes in the health of fish populations, 

affecting reproductive ability and fitness, especially in water bodies that are affected by episodic 

acidification. While there is evidence that ANC levels above 50 can confer additional protection 

from adverse ecological effects associated with aquatic acidification in some sensitive 

ecosystems, the expectation that such incremental protection from adverse effects will continue 

up to an ANC level of 100 is substantially reduced. The PA concludes that the above 

considerations support a focus on target ANC levels up to a level greater than 50 μeq/L but 

below 100 μeq/L, such as up to a level of 75 μeq/L.

In considering the available scientific evidence, as summarized here and discussed in 

more detail in the ISA and REA, in its review of the second draft PA, CASAC expressed the 

following views about the range of biological responses that corresponds to this range of ANC 

levels (i.e., 0-100 μeq/L):

There will likely be biological effects of acidification at higher ANC values 

within this range, and there are relatively insensitive organisms that are not 

impacted at ANC values at the low end of this range. Adverse effects of 

acidification on aquatic biota are fairly certain at the low end of this range of 
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ANC and incremental benefits of shifting waters to higher ANC become 

more uncertain at higher ANC levels. There is substantial confidence that 

there are adverse effects at ANC levels below 20 μeq/L, and reasonable 

confidence that there are adverse effects below 50 μeq/L. Levels of 50 

μeq/L and higher would provide additional protection, but the Panel has less 

confidence in the significance of the incremental benefits as the level 

increases above 50 μeq/L. (Russell and Samet, 2010b)

The PA concludes that the above considerations, including the views of CASAC, provide 

support for focusing on target ANC levels in the range of 20 to 75 μeq/L.

3. Consideration of Episodic Acidity 

As discussed in the PA, across the broad range of ANC values from 0 to 100 μeq/L, ANC 

affords protection against the likelihood of decreased pH (and associated increases in Al) during 

long or short periods. In general, the higher the ANC within this range, the lower the probability

of reaching low pH levels where direct effects such as increased fish mortality occur, as shown 

in Table 3-1 of the PA. Accordingly, greater protection would be achieved by target chronic 

ANC values set high enough to avoid pH depression to levels associated with elevated risk. 

The specific relationship between ANC and the probability of reaching pH levels of 

elevated risk varies by water body and fish species. The ANC levels below 20 μeq/L are 

generally associated with high probability of low pH, leading to death or loss of fitness of biota 

that are sensitive to acidification (US EPA, 2008, section 5.2.2.1; US EPA, 2009, section 

5.2.1.2). At these levels, during episodes of high acidifying deposition, brook trout populations 

may experience lethal effects. In addition, the diversity and distribution of zooplankton 

communities decline sharply at ANC levels below 20 μeq/L. Overall, there is little uncertainty

that significant effects on aquatic biota are occurring at ANC levels below 20 μeq/L. 

It is clear that at ANC levels approaching 0 μeq/L (Table II-1), there is significant 

impairment of sensitive aquatic ecosystems with almost complete loss of fish species. Avoiding 
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ANC levels approaching 0 μeq/L is particularly relevant to episodic spikes in acidity that occur 

during periods of rapid snow melt and during and after major precipitation events. Since the 

ANC range considered in the PA reflects average, long-term base flow values, it is appropriate to 

consider protecting against episodic drops in ANC values to a level as low as 0 μeq/L. Staddard 

et al. (2003) noted on average a 30 μeq/L depression of ANC between spring and summer time 

values, indicating the need to maintain higher base flow ANC levels to protect against ANC 

levels below 0 μeq/L. The above considerations do not provide support for a target chronic ANC 

level as low as 0 μeq/L for a standard that would protect against significant harm to aquatic 

ecosystems, including harm from episodic acidification. The PA concludes that these 

considerations also support a lower end of the range for consideration no lower than 20 μeq/L.

The CASAC agreed with this conclusion in its comments on the second draft PA (Russell 

and Samet, 2010b). The CASAC noted that “there are clear and marked biological effects at 

ANC values near 0 μeq/L, so this is probably not an appropriate target value” for the AAI. With 

regard to the likelihood of impairment of aquatic ecosystems due to episodic acidification, in 

terms of specific target levels for chronic ANC, CASAC expressed the following view:

Based on surface waters studied in the Northeast, decreases in ANC associated 

with snowmelt [are] approximately 50 μeq/L. Thus, based on these studies, a 

long term ANC target level of 75 μeq/L would generally guard against effects 

from episodic acidification down to a level of about 25 μeq/L. (Russell and 

Samet, 2010b)

4. Consideration of Ecosystem Response Time 

The PA notes that when considering a standard level to protect against aquatic 

acidification, it is appropriate to take into account both the time period to recovery as well as the 

potential for recovery in acid-sensitive ecoregions. Ecosystems become adversely impacted by 

acidifying deposition over long periods of time and have variable time frames and abilities to 
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recover from such perturbations. Modeling presented in the REA (US EPA, 2009, section 4.2.4) 

shows the estimated ANC values for Adirondack lakes and Shenandoah streams under preacidification conditions and indicates that for a small percentage of lakes and streams, natural 

ANC levels would have been below 50 μeq/L. Therefore, for these water bodies, reductions in 

acidifying deposition are not likely to achieve an ANC of 50 μeq/L or greater. Conversely, for 

some lakes and streams the level of perturbation from long periods of acidifying deposition has 

resulted in very low ANC values compared to estimated natural conditions. For such water 

bodies, the time to recovery would be largely dependent on future inputs of acidifying 

deposition.

Setting a standard level in terms of a target chronic ANC level is based on the long-term 

response of aquatic ecosystems. The time required for a water body to achieve the target ANC 

level -- given a decrease in ambient air concentrations of NOy and SOx and related acidifying 

deposition such that the critical load for a target ANC is not exceeded – is often decades if not 

centuries. In recognition of the potential public welfare benefits of achieving the target ANC in 

a shorter time frame, the concept of target loads had been developed. Target loads represent the 

depositional loading that is expected to achieve a particular level of the ecological indicator by a 

given time. For example, to achieve an ANC level of 20 μeq/L by 2030, it might be necessary to 

specify a higher target ANC level of, for example, 50 μeq/L, such that the depositional loading 

would be reduced more quickly than would occur if the depositional loading was based on 

achieving a target ANC level of 20 μeq/L as a long-term equilibrium level. In this example, the 

target ANC of 50 μeq/L would ultimately be realized many years later.

The above considerations have implications for selecting an appropriate standard level, in 

that the standard level affects not only the ultimate degree of protection that would be afforded 
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by the standard, but also the time frame in which such protection would be realized. However, 

the PA recognizes that there is a great deal of heterogeneity in response times among water 

bodies and that there is only very limited information from dynamic modeling that would help to 

quantify recovery time frames in areas across the country. As a consequence, quantification of a 

general relationship between critical loads associated with a specific long-term target ANC level 

and target loads associated with achieving the target ANC level within a specific time frame is 

not currently possible. Thus, while the time frame for recovery is an important consideration in 

selecting an appropriate range of levels to consider, the PA concludes that it can only be 

considered in a qualitative sense at this time.

5. Prior Examples of Target ANC Levels 

A number of regional organizations, states, and international organizations have 

developed critical load frameworks to protect against acidification of sensitive aquatic 

ecosystems. In considering the appropriate range of target ANC levels for consideration in this 

review, it is informative to evaluate the target ANC levels selected by these different 

organizations, as well as the rationale provided in support of the selected levels. Chapter 4 of the 

PA provides a detailed discussion of how critical loads have been developed and used in other 

contexts. Specific target values and their rationales are summarized below.

The UNECE has developed critical loads in support of international emissions reduction 

agreements. As noted in chapter 4 of the PA, critical loads were established to protect 95 percent 

of surface waters in Europe from an ANC less than 20 μeq/L based on protection of brown trout. 

Individual countries have set alternative ANC targets; for example, Norway targets an ANC of 

30 μeq/L based on protection of Atlantic salmon. Several states have established target ANC or 

pH values related to protection of lakes and streams from acidification. While recognizing that 
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some lakes in the Adirondacks will have a naturally low pH, the state of New York has 

established a target pH value of 6.5 for lakes that are not naturally below 6.5. As noted above, 

this level is associated with an ANC value that is likely to be between 20 and 50 μeq/L or 

possibly higher. New Hampshire and Vermont have set ANC targets of 60 μeq/L and 50 μeq/L, 

respectively. Tennessee has established site-specific target ANC values based on assessments of 

natural acidity, with a default value of 50 μeq/L when specific data are not available.

Taken together, these policy responses to concerns about ecological effects associated 

with aquatic acidification indicate that target ANC values between 20 and 60 μeq/L have been 

selected by states and other nations to provide protection of lakes and streams in some of the 

more sensitive aquatic ecosystems.

6. Consideration of Public Welfare Benefits

The point at which effects on public welfare become adverse is not defined in the CAA. 

Characterizing a known or anticipated adverse effect to public welfare is an important 

component of developing any secondary NAAQS. According to the CAA, welfare effects 

include:

...effects on soils, water, crops, vegetation, manmade materials, animals, 

wildlife, weather, visibility, and climate, damage to and deterioration of 

property, and hazards to transportation, as well as effect on economic values and 

on personal comfort and well-being, whether caused by transformation, 

conversion, or combination with other air pollutants. (CAA, section 302(h)).

Consideration of adversity to public welfare in the context of the secondary NAAQS for oxides 

of nitrogen and sulfur can be informed by information about losses in ecosystem services 

associated with acidifying deposition and the potential economic value of those losses, as 

summarized above in section II.C and discussed more fully in chapter 4 of the PA. 
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Ecosystem service losses at alternative ANC levels are difficult to enumerate. However, 

in general there are categories of ecosystem services, discussed in chapter 4 of the PA, that are 

related to the specific ecosystem damages expected to occur at alternative ANC levels. Losses in 

fish populations due to very low ANC (below 20 μeq/L) are likely associated with significant 

losses in value for recreational and subsistence fishers. Many acid sensitive lakes are located in 

areas with high levels of recreational fishing activity. For example, in the northeastern U.S., 

where nearly 8 percent of lakes are considered acidic, more than 9 percent of adults participate in 

freshwater fishing, with an estimated value of approximately $5 billion in 2006. This suggests 

that improvements in lake fish populations may be associated with significant recreational 

fishing value.

As discussed in the PA, inland surface waters also provide cultural services such as 

aesthetic and existence value and educational services. To the extent that piscivorous birds and 

other wildlife are harmed by the absence of fish in these waters, hunting and birdwatching 

activities are likely to be adversely affected. A case study of the value to New York residents of 

improving the health of lakes in the Adirondacks found significant willingness to pay for those 

improvements. When scaled to evaluate the improvement in lake health from achieving ANC 

values of either 20 or 50 μeq/L, the study implies benefits to the New York population roughly 

on the order of $300 - 900 million per year (in constant 2007$). The survey administered in this 

study recognized that participants were thinking about the full range of services provided by the 

lakes in question – not just the recreational fishing services. Therefore the estimates of 

willingness to pay include resident’s benefits for potential hunting and birdwatching activities 

and other ancillary services. These results are just for New York populations. The PA concludes 

that if similar benefits exist for improvements in other acid sensitive lakes, the economic value to 
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U.S. populations could be very substantial, suggesting that, at least by one measure of impact on 

public welfare, impacts associated with ANC less than 50 μeq/L may be adverse to public 

welfare.

7. Summary of Alternative Levels

Based on all the above considerations, the PA concludes that consideration should be 

given to a range of standard levels from 20 to 75 μeq/L. The available evidence indicates that 

target ANC levels below 20 μeq/L would be inadequate to protect against substantial ecological 

effects and potential catastrophic loss of ecosystem function in some sensitive aquatic 

ecosystems. While ecological effects occur at ANC levels below 50 μeq/L in some sensitive 

ecosystems, the degree and nature of those effects are less significant than at levels below 20 

μeq/L. Levels at and above 50 μeq/L would be expected to provide additional protection, 

although uncertainties regarding the potential for additional protection from adverse ecological 

effects are much larger for target ANC levels above about 75 μeq/L, as effects are generally 

appreciably less sensitive to changes in ANC at such higher levels.

In reaching this conclusion in the PA, consideration was given to the extent to which a 

target ANC level within this range would protect against episodic as well as long-term ecological 

effects. Levels in the mid- to upper part of this range would be expected to provide greater 

protection against short-term, episodic peaks in aquatic acidification, while lower levels within 

this range would give more weight to protection from long-term rather than episodic 

acidification. Similarly, levels in the mid- to upper part of this range would be expected to result 

in shorter time periods for recovery given the lag in ecosystem response in some sensitive 

ecosystems relative to levels in the lower part of this range. The PA also notes that this range 
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encompasses target ANC values that have been established by various States and regional and 

international organizations to protect against acidification of aquatic ecosystems.

The PA recognizes that the level of the standard together with the other elements of the 

standard, including the ambient air indicators, averaging time, and form, determine the overall 

protectiveness of the standard. Thus, consideration of a standard level should reflect the 

strengths and limitations of the evidence and assessments as well as the inherent uncertainties in 

the development of each of the elements of the standard. The implications of considering 

alternative standards, defined in terms of alternative combinations of levels and percentile values 

that are a critical component of factor F1 in the form of the standard, are discussed below in 

section III.E. Key uncertainties in the various components of the standard are summarized and 

considered below in section III.F. 

E. Combined Alternative Levels and Forms

To provide some perspective on the implications of various alternative multi-pollutant, 

AAI-based standards, the PA presented the number of acid-sensitive ecoregions that would likely

not meet various sets of alternative standards. The alternative standards considered were based 

on combinations of alternative target ANC levels, within the range of 20 to 75 μeq/L, and 

alternative forms, characterized by alternative representative percentiles within the range of the 

70th to 90th percentile. These alternative standards are also defined in terms of the other elements 

of the standard: ambient air indicators NOy and SOx, discussed above in section III.A; other 

elements of the form of the standard, including ecoregion-specific values for factors F1 through 

F4 in the AAI equation, discussed above in section III.B.5; and an annual averaging time for 

NOy and SOx, discussed above in section III.C. With regard to the averaging time, the 

assessment did not consider multi-year averaging of the calculated annual AAI values due to data 
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limitations, including, for example, the lack of CMAQ modeling for multiple consecutive years. 

In this assessment, we characterize an ecoregion as likely not meeting a given alternative 

standard if the calculated AAI value is less than the target ANC level of the standard, 

recognizing that higher AAI values are more protective than lower values.

The results of this assessment are presented in Table 7-1a-d in the PA for a subset of 

ecoregions including those characterized as acid sensitive. Calculated annual AAI values at the 

ecoregion level are shown for each alternative standard considered. Based on these AAI values, 

Table 7-2 in the PA summarizes the number of acid-sensitive ecoregions that would likely not 

meet each of the alternative standards considered. 14 Calculated AAI values for all ecoregions 

categorized as relatively non-acid sensitive are shown in Table D-5 in Appendix D of the PA. In 

all cases, these relatively non-acid sensitive ecoregions were estimated to meet all of the 

alternative standards considered in this assessment.

As described above, the AAI values presented in Table 7-1a-d of the PA are based in part 

on data from 2005 CMAQ model simulations, which was used to generate values for F2 through 

F4 in the AAI equation, as well as to estimate annual average ambient air concentrations of NOy

and SOx that reflect recent air quality in the absence of currently available monitored 

concentrations in sensitive ecoregions across the country. Water quality and hydrology data 

from water bodies within each ecoregion were also used in calculating the AAI values. Such 

data were initially used to calculate critical loads for each water body with sufficient data within 

an ecoregion so as to identify the nth percentile critical load representative of the ecoregion used 

in calculating the F1 factor for the ecoregion. As expected, the number of ecoregions that likely 

 

14 Tables 7-1a-d and 7-2 in the PA present assessment results for 29 ecoregions that had been initially characterized 

as acid sensitive. Subsequently, based on a broader set of criteria used to characterize ecoregions as acid sensitive, 

as discussed above in section III.B.5.a, the set of ecoregions characterized as acid sensitive was narrowed to include 

22 ecoregions.
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would not meet alternative standards increases with increasing percentile values and target ANC 

levels (US EPA, 2011, Table 7-2). Out of 22 acid-sensitive ecoregions, the number of 

ecoregions that would likely not meet the alternative standards ranges from 22 for the most 

protective alternative standard considered (75 μeq/L, 90th percentile) to 4 for the least protective 

alternative standard (20 μeq/L, 70th percentile). It is apparent that both the percentile and the 

level chosen have a strong influence, over the ranges considered, in determining the number of 

areas that would likely not meet this set of alternative standards. 

The PA observes that there is one grouping of these acid-sensitive ecoregions that would 

likely not meet almost all combinations of level and form under consideration (US EPA, 2011,

Table 7-2 and Appendix D). This group is made up of southern Appalachian mountain areas, 

including North Central Appalachians, 5.3.3; Ridge and Valley, 8.4.1; Central Appalachians, 

8.4.2; Blue Ridge, 8.4.4; and Southwestern Appalachians, 8.4.9. In addition, these ecoregions 

exhibit the highest amounts of exceedance relative to alternative standards.

The Northern Appalachian and Atlantic Maritime Highlands (5.3.1), which includes the 

Adirondacks, and the Northern Lakes and Forests (5.2.1) of the upper midwest exhibit similar 

patterns with respect to in the role of level and percentile in identifying regions not likely to meet 

alternative standards, although there are considerably fewer cases compared to the regions in the 

Appalachians.

In the mountainous west, the Sierra Nevada (6.2.12), Idaho Batholith (6.2.15) and the 

Cascades (6.2.7) ecoregions likely would not meet alternative standards in fewer cases relative to 

eastern regions, with the Sierra Nevada ecoregion exhibiting relatively greater sensitivity

compared to all western regions. Only in the upper part of the ranges of level and percentile do 

regions in the northern and central Rockies likely not meet alternative standards.
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In considering these findings, the PA observes that the standard as defined by the AAI 

behaves in an intuitively logical manner. That is, an increase in ecoregions likely not to meet the 

standard is associated with higher alternative levels and percentiles, both of which contribute to a 

lower regionally representative critical load. Moreover, the areas of known adverse aquatic 

acidification effects are identified, mostly in high elevation regions or in the northern latitudes --

the Adirondacks, Shenandoahs, northern midwest lakes and the mountainous west. These results 

reflect the first application of a nationwide model that integrates water quality and atmospheric 

processes at a national scale and provides findings that are consistent with our basic 

understanding of the extent of aquatic acidification across the U.S. What is particularly 

noteworthy is that this model is not initialized with a starting ANC based on water quality data, 

which likely would result in a reproduction of water quality observations. Rather, this standard 

reflects the potential of the changes in atmospheric concentrations of NOy and SOx to induce 

long-term sustained changes in surface water systems. The PA notes that the fact that the 

patterns of adversity based on applying this standard are commensurate with what is observed in 

surface water systems provides confidence in the basic underlying formulation of the standard.

The PA notes that the Appalachian mountain regions merit further inspection as they

stand out as areas with the largest relative exceedances from a national perspective. Water 

quality data from these regions as well as an emissions sensitivity CMAQ simulation were 

considered to better understand the simulated behavior of these regions. The maps and tables in 

appendix D of the PA include paired comparisons of the CMAQ 2005 and emissions sensitivity

simulations. The emissions sensitivity simulation reflects domain-wide reductions in NOy and 

SOx emissions of 48 percent and 42 percent, respectively, relative to 2005 base year emissions. 

The PA assumes that this emissions sensitivity simulation is indicative of future conditions. 
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The emissions sensitivity results project that many of the regions that likely would not 

meet the alternative standards based on recent air quality, especially at alternative levels of 20 

and 35 μeq/L, would likely meet such standards in the future year scenario for the Appalachian 

mountain regions. It is apparent that the AAI calculations are especially sensitive to changes in 

SOx emissions as the Appalachian regions have the highest SOx concentrations and deposition 

rates (US EPA, 2011,section 2), and the AAI equation responds as expected to modeled 

reductions in SOx. The emissions sensitivity scenario is a prospective application of the 

standard, in the sense that rules derived from the air quality management process result in 

reductions of NOy and SOx emissions. Expected emission changes over the next two decades 

should be far greater than the 42 percent and 48 percent, respectively, SOx and NOy reductions 

used in this analysis, with a consequent further reduction in areas that would likely not meet 

alternative standards.

The Appalachian mountain regions generally have low DOC levels, average runoff rates, 

moderately low base cation supply and highly elevated sulfate concentrations. Collectively, 

those attributes do not suggest naturally acidic conditions as the availability of anthropogenic 

contributions of mineral acids is likely responsible for observed low ANC values in those 

regions.

The PA notes the Sierra Nevada region as an interesting case study, as it has some of the 

lowest critical load values nationally (US EPA, 2011, Table D-3). Water quality data indicate 

extremely low sulfate, as expected given the relatively low SO2 emissions in the western U.S. 

Extremely low base cation supply and low Neco, which mitigate the effect of nitrogen 

deposition, explain the low critical load values. Low Neco values appear to associate well with 

high elevation western U.S. regions, perhaps reflecting the more arid and reduced vegetation 
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density relative to eastern U.S. regions. The proximity to high level nitrogen emissions 

combined with very low base cation supply explains the cases where the Sierra region likely 

does not meet alternative standards. Because Neco values are low in the Sierras, the system 

responds effectively to reductions of NOx emissions, as illustrated in the maps and tables of 

Appendix D of the PA. Although Neco affords protection from the acidifying effects of nitrogen 

deposition, the availability of excessive nitrogen neutralization capacity also means that 

reductions in nitrogen are not as effective as reductions in SOx in reducing the calculated AAI.

In reviewing these results, the PA observes that the analysis of the alternative 

combinations of level and form presented provide context for considering the impact of different 

standards. Since the AAI equation has been newly developed in the PA, these examples of 

estimated exceedances help to address the question of whether the AAI equation responds in a 

reasonable manner with regard to identifying areas of concern and to prospective changes in 

atmospheric conditions likely to result from future emissions reduction strategies. The PA

concludes that the behavior of the AAI calculations is both reasonable and explainable, which 

the PA concludes serves to increase confidence in considering a standard defined in terms of the 

AAI.

F. Characterization of Uncertainties

This section summarizes discussions of the results of analyses and assessments, presented 

more fully in the PA (US EPA, 2011, section 7.6 and Appendices F and G), intended to address 

the relative confidence associated with the linked atmospheric-ecological effects system 

described above. An overview of uncertainties is presented in the context of the major structural 

components underlying the standard, as well as with regard to areas of relatively high 

uncertainty. The section closes with a discussion of data gaps and uncertainties associated with 
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the use of ecological and atmospheric modeling to specify the factors in the AAI equation, which 

can be used to guide future field programs and longer-term research efforts. 

1. Overview of Uncertainty

As discussed in the PA (US EPA, 2011,Table 7-3), there is relatively low uncertainty 

with regard to the conceptual formulation of the overall structure of the AAI-based standard that

incorporates the major associations linking biological effects to air concentrations. Based on the 

strength of the evidence that links species richness and mortality to water quality, the 

associations are strongly causal and without any obvious confounding influence. The strong 

association between the ecosystem indicator (ANC) and the causative water chemistry species 

(dissolved aluminum and hydrogen ion) reinforces the confidence in the linkage between 

deposition of nitrogen and sulfur and effects. This strong association between ANC and effects 

is supported by a sound mechanistic foundation between deposition and ANC. The same 

mechanistic strength holds true for the relationship between ambient air levels of nitrogen and 

sulfur and deposition, which completes the linkage from ambient air indicators through 

deposition to ecological effects. 

There are relatively higher uncertainties, however, in considering specific elements

within the structure of an AAI-based standard, including the deposition of SOx, NOy, and NHx as 

well as the critical load-related component, each of which can vary within and across ecoregions. 

Overall system uncertainty relates not just to the uncertainty in each such element, but also to the 

combined uncertainties that result from linking these elements together within the AAI-based 

structure. Some of these elements – including, for example, dry deposition, pre-industrial base 

cation production, and reduced nitrogen deposition – are estimated with less confidence than 

other elements (US EPA, 2011, Table 7.3). The uncertainties associated with all of these 
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elements, and the combination of these elements through the AAI equation, are discussed below

and in the following sections related to measured data gaps and modeled processes for both air 

quality and water quality.

The lack of observed dry deposition data is constrained by resources and the lack of 

efficient measurement technologies. Progress in reducing uncertainties in dry deposition will 

depend on improved atmospheric concentration data and direct deposition flux measurements of 

the relevant suite of NOy and SOx species. 

Pre-industrial base cation productivity by definition is not observable. Contemporary

observations and inter-model comparisons are useful tools that would help reduce the uncertainty

in estimates of preindustrial base cation productivity used in the AAI equation. In characterizing 

contemporary base cation flux using basic water quality measurements (i.e., major anion and 

cation species as defined in equation 2.11 in the PA), it is reasonable to assume that a major 

component of contemporary base cation flux is associated with pre-industrial weathering rates. 

To the extent that multiple models converge on similar solutions, greater confidence in 

estimating pre-industrial base cation production would be achieved.

Characterization of NHx deposition has been evolving over the last decade. The 

relatively high uncertainty in characterizing NHx deposition is due to both the lack of field 

measurements and the inherent complexity of characterizing NHx with respect to source 

emissions and dry deposition. Because ammonia emissions are generated through a combination 

of man-made and biological activities, and ammonia is semi-volatile, the ability to characterize 

spatial and temporal distributions of NHx concentrations and deposition patterns is challenging. 

While direct measurement of NHx deposition is resource intensive because of the diffuse nature 

of sources (i.e., area-wide and non-point sources), there have been more frequent deposition flux 
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studies, relative to other nitrogen species, that enable the estimation of both emissions and dry

deposition. Also, while ammonia has a relatively high deposition velocity and traditionally was

thought to deposit close to the emissions release areas, the semi-volatile nature of ammonia 

results in re-entrainment back into the lower boundary layer resulting in a more dispersed 

concentration pattern exhibiting transport type characteristics similar to longer lived atmospheric 

species. These inherent complexities in source characterization and ambient concentration 

patterns raise the uncertainty level of NHx in general. However, the PA notes that progress is 

being made in measuring ammonia with cost efficient samplers and anticipates the gradual 

evolution of a spatially robust ammonia sampling network that would help support analyses to 

reduce underlying uncertainties in NHx deposition. Also, from an aquatic acidification 

perspective, NHx is not as important a driver as NOy and SOx in the mountainous areas in the

eastern U.S. However, the relative importance of NHx is likely to increase over time, in light of

air quality rules in place designed to reduce emissions of NOy and SOx.

2. Uncertainties Associated with Data Gaps

In summarizing uncertainties with respect to available measurement data and the use of 

ecological and atmospheric models, the PA indentified data gaps and model uncertainties in 

relative terms by comparing, for example, the relative richness of data between geographic areas 

or environmental media. With regard to relevant air quality measurements, the PA notes that 

such measurements are relatively sparse in the western U.S. While the spatial extent of 

CASTNET coverage has gradually incorporated western U.S. locations with support from the 

NPS, the relative density of monitoring sites is much less than that in the eastern U.S. This 

relative disparity in spatial density of monitors is exacerbated as air quality patterns in the 

mountainous west generally exhibit greater spatial heterogeneity due to dramatic elevation 
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gradients that impact meteorology and air mass flow patterns. Similarly, water quality data 

coverage is far more comprehensive in the eastern U.S. relative to the west

Measurements of NOy notably are lacking in both eastern and western acid-sensitive 

ecoregions. This adds uncertainty to the use of the AAI equation as the lack of NOy data limits 

efforts to evaluate air quality modeling of NOy that is the basis for quantifying factor F3 in the 

AAI equation. The lack of NOy measurements also limits efforts to characterize the variability

and representativeness of modeled NOy concentrations within and across ecoregions. Currently, 

the Agency’s ability to define the protection likely to be afforded by alternative standards (in 

terms of alternative levels and percentiles) is compromised by the lack of a full set of ambient air 

quality indicator measurements, notably including NOy, throughout sensitive ecoregions across 

the U.S. 

Further, obtaining measurement of the dominant species that comprise NOy (HNO3, true 

NO2, NO, p–NO3, and PAN) would be useful to evaluate performance of NOy samplers. Beyond 

the more well known dominant components of NOy, research efforts would be needed to 

characterize total reactive nitrogen that may include significant amounts of organically-bound 

nitrogen (beyond PAN) which is poorly understood with regard to emission sources and 

concentration levels.

Field measurements of NHx have been extremely limited, but have begun to be enhanced 

through the NADP’s passive ammonia network (AMoN). The AMoN measures ammonia at 

over 50 sites, with more than 35 at CASTNET locations. Enhanced spatial coverage of reduced 

nitrogen measurements, particularly to understand within and across ecoregion variability, and 

the inclusion of some continuous observations would provide a better understanding of the 
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uncertainty in the F2 factor in the AAI equation and of the representativeness of modeled NHx

deposition within and across ecoregions.

With regard to water quality data, the PA notes that such data are typically limited 

relative to air quality data sets, and are also relatively sparse in the western U.S. The 

TIME/LTM water quality sampling program in the eastern U.S. (as described in chapter 2 of the 

PA) is an appropriate complement to national air monitoring programs as it affords consistency

across water bodies in terms of sampling frequency and analysis protocols. Consideration should 

be given to extending the TIME/LTM design to all acid sensitive ecoregions, with priority for

areas in the western mountains that are data limited and showing initial signs of adversity 

particularly with respect to aquatic acidification. The lack of a regulatory requirement for 

TIME/LTM often jeopardizes funding support of this resource that is especially valuable and 

cost effective. While there are several state and local agency water quality data bases, it is 

unclear the extent to which differences in sampling, chemical analysis and reporting protocols 

would impact the use of such data for the purpose of better understanding the degree of 

protectiveness that would be afforded by an AAI-based standard within sensitive ecoregions 

across the country. In addition, our understanding of water quality in Alaska and Hawaii and the 

acid sensitivity of their ecoregions is particularly limited. 

Water quality data and modeling support the standard setting process. As more water 

bodies are sampled, the critical load data bases would expand, enabling clearer delineation of 

ecoregion representative critical loads in terms of the nth percentile. This would provide more 

refined characterization of the degree of protection afforded by a given standard. Longer term, 

the availability of water quality trend data (annual to monthly sampled) would support 

accountability assessments that examine if an ecoregion’s response to air management efforts is 
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as predicted by earlier model forecasting. The most obvious example is the long-term response 

of water quality ANC change to changes in calculated AAI, deposition, ambient NOy and SOx

concentrations, and emissions. In addition, water quality trends data provide a basis for 

evaluating and improving the parameterizations of processes in critical load models applied at 

the ecoregion scale related to nitrogen retention and base cation supply. A better understanding

of soil processes, especially in the southern Appalachians, would enhance efforts to examine the 

variability within ecoregions of the soil-based adsorption and exchange processes which 

moderate the supply of major cations and anions to surface waters and strongly influence the 

response of surface water ANC to changes in deposition of nitrogen and sulfur.

3. Uncertainties in Modeled Processes

As discussed in the PA, from an uncertainty perspective, gaps in field measurement data 

are related to uncertainties in modeled processes and in the specific application of such models. 

As noted above, processes that are embodied in an AAI-based standard are modeled using the 

CMAQ atmospheric model and steady state ecological models. These models are characterized 

in the ISA as being well established and they have undergone extensive peer review. 

Nonetheless, the application of these models for purposes of specifying the factors in the AAI 

equation, on an ecoregion scale, is a new application that introduces uncertainties, as noted 

below, especially in areas with limited observational data that can be used to evaluate this 

specific application. Understanding uncertainties in relevant modeled process thus involves 

consideration of the uncertainties associated with applying each model as well as the 

combination of these uncertainties as the models are applied in combination within the AAI 

framework.
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With regard to the application of CMAQ for purposes of use in an AAI-based standard, 

the modeling of dry deposition has been identified as having a relatively high degree of 

uncertainty. Due to a combination of system complexity and resource constraints, there is no 

routine observational basis for directly comparing modeled dry deposition and measurements. 

Periodic dry deposition flux experiments covering a variety of vegetation, surfaces and 

meteorology across seasons would enable a more robust evaluation of modeled deposition of 

nitrogen and sulfur. Given the difficulty in acquiring dry deposition observations, it becomes 

especially important to evaluate the model’s ability to capture temporal and spatial ambient air 

patterns of individual nitrogen and sulfur species which are used to drive dry deposition 

calculations in models. For example, reducing a generally acknowledged positive bias in modelpredicted SO2 relative to observations is especially relevant to the AAI-based standard, as SO2

deposition is a dominant contributor to total acidifying deposition in the eastern U.S. With 

respect to oxidized nitrogen, observations of individual NOy species are important as air quality 

models calculate the individual deposition of each species. The modeled transference ratios, 

TNOy and TSOx used in factors F3 and F4 rely on CMAQ’s ability to characterize both deposition 

and concentration. Consequently, a better understanding of the variability of these factors 

within and across ecoregions could be achieved by improved availability of measured ambient 

concentrations and deposition observations.

Steady state biogeochemical ecosystem modeling is used to develop critical load 

estimates that are incorporated in the AAI equation through factor F1. Consequently, the PA

notes that an estimate of the temporal response of surface water ANC to deposition and air 

concentration changes is not directly available. Lacking a predicted temporal response impairs 

the ability to conduct accountability assessments down to the effects level. Accountability
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assessments would examine the response of each step in the emissions source through air 

concentration – deposition -- surface water quality – biota continuum. The steady state 

assumption at the ecosystem level does not impair accountability assessments through the air 

concentration/deposition range of that continuum. However, in using steady state ecosystem 

modeling, several assumptions are made relative to the long-term importance of processes 

related to soil adsorption of major ions and ecosystem nitrogen dynamics. Because these models 

often were developed and applied in glaciated areas with relatively thin and organically rich 

soils, their applicability is relatively more uncertain in areas such as those in the non-glaciated 

clay-based soil regions of the central Appalachians. Consequently, it is desirable to develop the 

information bases to drive simple dynamic ecosystem models that incorporate more detailed 

treatment of subsurface processes, such as adsorption and exchange processes and sulfate 

absorption.

4. Applying Knowledge of Uncertainties

An understanding of the relative uncertainties in a system assists in setting priorities for 

data collection efforts and research, with the expectation that such efforts would reduce 

uncertainties over time and afford greater confidence in applications of an AAI-based standard. 

Because of the uniquely wide breadth of pollutants and environmental media addressed by an

AAI-based multi-pollutant standard, there are a wide range of uncertainties that are important to 

consider relative to single pollutant standards that typically address only direct effects of ambient 

air exposures. For an AAI-based standard, a reduction of the uncertainties across the various 

modeled processes at the ecoregion scale would lead to greater confidence in the degree of 

protection afforded by the standard. 
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The PA notes that there is generally low uncertainty with regard to the conceptual 

development and related major components of this standard. In recognizing the scientific 

soundness of the basic structure of this standard, the PA notes that future efforts would be

appropriately directed at expanding the availability of relevant data for ecoregion-specific 

evaluation and application of the relevant modeling of ecological and atmospheric processes, as 

identified above. Such efforts would further support consideration of an AAI-based standard and 

would guide field studies and analyses designed to improve the longer-term confidence in such a

standard.

G. CASAC Advice

The CASAC has advised EPA concerning the ISA, the REA, and the PA. The CASAC 

has endorsed EPA’s interpretation of the science embodied in the ISA and the assessment 

approaches and conclusions incorporated in the REA.

Most recently, CASAC has considered the information in the final PA in providing its 

recommendations on the review of the new multi-pollutant standard developed in that document 

and discussed above (Russell and Samet, 2011a). In so doing, CASAC has expressed general 

support for the conceptual framework of the standard based on the underlying scientific 

information, as well as for the conclusions in the PA with regard to indicators, form, averaging 

time, and level of the standard that are appropriate for consideration by the Agency in reaching 

decisions on the review of the secondary NAAQS for oxides of nitrogen and sulfur:

The final Policy Assessment clearly sets out the basis for the recommended ranges 

for each of the four elements (indicator, averaging time, level and form) of a 

potential NAAQS that uses ambient air indicators to address the combined effects 

of oxides of nitrogen and oxides of sulfur on aquatic ecosystems, primarily 

streams and lakes. As requested in our previous letters, the Policy Assessment also 

describes the implications of choosing specific combinations of elements and 

provides numerous maps and tabular estimates of the spatial extent and degree of 
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severity of NAAQS exceedances expected to result from possible combinations of 

the elements of the standard.

We believe this final PA is appropriate for use in determining a secondary

standard to help protect aquatic ecosystems from acidifying deposition of oxides 

of sulfur and nitrogen. EPA staff has done a commendable job developing the 

innovative Aquatic Acidification Index (AAI), which provides a framework for a 

national standard based on ambient concentrations that also takes into account 

regional differences in sensitivities of ecosystems across the country to effects of 

acidifying deposition. (Russell and Samet, 2011a)

The CASAC also recommended that as EPA moves forward in the regulatory process 

“some attention should be given to our residual concern that the available data may reflect the 

more sensitive water bodies and thus, the selection of percentiles of waterbodies to be protected 

could be conservatively biased” (Russell and Samet, 2011a). In addition, CASAC found some 

improvements could be made to the uncertainty analysis, as noted below. With respect to 

indicators, CASAC supports the use of SOx and NOy as ambient air indicators (discussed above 

in section III.A) and ANC as the ecological indicator (discussed above in section III.B.1):

The use of NOy and SOx as atmospheric indicators of oxides of nitrogen and sulfur 

atmospheric concentrations is well justified. The use in the AAI of NOy and SOx as atmospheric 

indicators of oxides of nitrogen and sulfur concentrations is useful and corresponds with other 

efforts by EPA. As we have stated previously, CASAC also agrees that ANC is the most 

appropriate ecological indicator of aquatic ecosystem response and resiliency to acidification 

(Russell and Samet, 2011a).

With respect to the form of the standard (discussed above in section III.B), CASAC 

stated the following:

EPA has developed the AAI, an innovative “form” of the NAAQS itself that 

incorporates the multi-pollutant, multi-media, environmentally modified, 

geographically variable nature of SOx/NOy deposition-related aquatic 

acidification effects. With the caveats noted below, CASAC believes that this 
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form of the NAAQS as described in the final Policy Assessment is consistent with 

and directly reflective of current scientific understanding of effects of acidifying 

deposition on aquatic ecosystems. (Russell and Samet, 2011a)

CASAC agrees that the spatial components of the form in the Policy Assessment 

are reasonable and that use of Omernick’s ecoregions (Level III) is appropriate for 

a secondary NAAQs intended to protect the aquatic environment from 

acidification . . . (Russell and Samet, 2011a)

The “caveats” noted by CASAC include a recognition of the importance of continuing to 

evaluate the performance of the CMAQ and ecological models to account for model 

uncertainties and to make the model-dependent factors in the AAI more transparent. In addition, 

CASAC noted that the role of DOC and its effects on ANC would benefit from further 

refinement and clarification (Russell and Samet, 2011a). While CASAC expressed the view that 

the “division of ecoregions into ‘sensitive’ and ‘non-sensitive’ subsets, with a more protective 

percentile applied to the sensitive areas, also seems reasonable” (Russell and Samet, 2011a),

CASAC also noted that there was the need for greater clarity in specifying how appropriate 

screening criteria would be applied in assigning ecoregions to these categories. Further, CASAC 

identified potential biases in critical load calculations and in the regional representativeness of 

available water chemistry data, leading to the observation that a given percentile of the 

distribution of estimated critical loads may be protective of a higher percentage of surface waters 

in some regions (Russell and Samet, 2011a).

With respect to averaging time (discussed above in section III.C), CASAC stated the 

following:

Considering the cumulative nature of the long-term adverse ecological effects and 

the year-to-year variability of atmospheric conditions (mainly in the amount of 

precipitation), CASAC concurs with EPA that an averaging time of three to five 

years for the AAI parameters is appropriate. A longer averaging time would mask 

possible trends of AAI, while a shorter averaging time would make the AAI being 
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more influenced by the conditions of the particular years selected. (Russell and 

Samet, 2011a)

With respect to level as well as the combination of level and form as they are presented 

as alternative standards (discussed above in sections III.D-E), CASAC stated the following:

CASAC agrees with EPA staff’s recommendation that the “level” of the 

alternative AAI standards should be within the range of 20 and 75 μeq/L. We also 

recognize that both the “level” and the form of any AAI standard are so closely

linked in their effectiveness that these two elements should be considered 

together. (Russell and Samet, 2011a)

When considered in isolation, it is difficult to evaluate the logic or implications 

of selecting from percentiles (70th to 90th) of the distribution of estimated critical 

loads for lakes in sensitive ecoregions to determine an acceptable amount of 

deposition for a given ecoregion. However, when these percentile ranges are 

combined with alternative levels within the staff-recommended ANC range of 20 

to 75 microequivalents per liter (μeq/L), the results using the AAI point to the 

ecoregions across the country that would be expected to require additional 

protection from acidifying deposition. Reasonable choices were made in 

developing the form. The number of acid sensitive regions not likely to meet the 

standard will be affected both by choice of ANC level and the percentile of the 

distribution of critical loads for lakes to meet alternative ANC levels in each 

region. These combined recommendations provide the Administrator with a 

broad but reasonable range of minimally to substantially protective options for the 

standard. (Russell and Samet, 2011a)

CASAC also commented on EPA’s uncertainty analysis, and provided advice on areas 

requiring further clarification in the proposed rule and future research. The CASAC found it 

“difficult to judge the adequacy of the uncertainty analysis performed by EPA because of lack of 

details on data inputs and the methodology used, and lack of clarity in presentation” (Russell and 

Samet, 2011a). In particular, CASAC identified the need for more thorough model evaluations 

of critical load and atmospheric modeling, recognizing the important role of models as they are 

incorporated in the form of the standard. In light of the innovative nature of the standard 

developed in the PA, CASAC identified “a number of areas that should be the focus of further 

research” (Russell and Samet, 2011a). While CASAC recognized that EPA staff was able to 
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address some of the issues in the PA, they also noted areas “that would benefit from further study

or consideration in potential revisions or modifications to the form of the standard.” Such 

research areas include “sulfur retention and mobilization in the soils, aluminum availability, soil 

versus water acidification and ecosystem recovery times.” Further, CASAC encouraged future 

efforts to monitor individual ambient nitrogen species, which would help inform further CMAQ 

evaluations and the specification of model-derived elements in the AAI equation (Russell and 

Samet, 2011a).

H. Administrator’s Proposed Conclusions

Having concluded that the existing NO2 and SO2 secondary standards are neither 

sufficiently protective nor appropriate to address deposition-related effects associated with 

oxides of nitrogen and sulfur (section II.D above), the Administrator has considered whether it is 

appropriate at this time to set a new multi-pollutant standard for that purpose, with a structure 

that would better reflect the available science regarding acid deposition. In considering this, she 

recognizes that such an appropriate standard, for purposes of section 109(b) and (d) of the 

CAA,15, must in her judgment be requisite to protect public welfare, such that it would be neither 

more nor less stringent that necessary for that purpose. In particular, she has focused on the new 

standard developed in the PA and reviewed by CASAC, as discussed above. In so doing, the 

Administrator first considered the extent to which there is a scientific basis for development of 

such a standard, specifically with regard to a standard that would provide protection from 

deposition-related aquatic acidification in sensitive aquatic ecosystems in areas across the 

country. As discussed above, the Administrator notes that the ISA concludes that the available

scientific evidence is sufficient to infer a causal relationship between acidifying deposition of 

 

15 Section 109(d)(1) requires that “...the Administrator shall complete a thorough review...and shall make such 

revisions in such criteria and standards and promulgate such new standards as may be appropriate 

under...subsection 109(b) of this section.” [emphasis added]
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nitrogen and sulfur in aquatic ecosystems, and that the deposition of oxides of nitrogen and 

sulfur both cause such acidification under current conditions in the U.S. Further, the ISA

concludes that there are well-established water quality and biological indicators of aquatic 

acidification as well as well-established models that address deposition, water quality, and 

effects on ecosystem biota, and that ecosystem sensitivity to acidification varies across the 

country according to present and historic nitrogen and sulfur deposition as well as geologic, soil, 

vegetative, and hydrologic factors. Based on these considerations, the Administrator agrees with 

the conclusion in the PA, and supported by CASAC, that there is a strong scientific basis for 

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

The Administrator also recognizes that the conceptual framework for an ecologically

relevant, multi-pollutant standard, which was initially explored in the REA and further developed 

in the PA, builds on the information in the ISA. She notes that the structure of the standard 

addresses the combined effects of deposition from oxides of nitrogen and sulfur by 

characterizing the linkages between ambient concentrations, deposition, and aquatic 

acidification, and that the structure of the standard takes into account relevant variations in these 

linkages across the country. She recognizes that while the standard is innovative and unique, the 

structure of the standard is well grounded in the science underlying the relationships between 

ambient concentrations of oxides of nitrogen and sulfur and the aquatic acidification related to 

deposition of nitrogen and sulfur associated with such ambient concentrations.

While the Administrator recognizes the strong scientific foundation for the structure of an 

AAI-based standard, she also recognizes that the standard depends on atmospheric and 

ecological modeling, based on appropriate data, to specify the terms of an equation that 

incorporates the linkages between ambient concentrations, deposition, and aquatic acidification. 
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This equation, which defines an aquatic acidification index (AAI), has the effect of translating 

spatially variable ecological effects into a potential national standard. With respect to 

establishing the specific terms of this equation, there are a number of inherent uncertainties and 

complexities that are relevant to the question of whether it is appropriate under section 109 to set 

a specific AAI-based standard at this time, recognizing that such a standard must be requisite to 

protect public welfare without being either more or less stringent than necessary for this purpose. 

As discussed above, these uncertainties and complexities generally relate not to the structure of 

the standard, but to the quantification of the various elements of the standard, such as the F 

factors discussed earlier in this section and their representativeness at an ecoregion scale. These 

uncertainties and complexities currently limit efforts to characterize the degree of protectiveness 

that would be afforded by such a standard, within the ranges of levels and forms identified in the 

PA, and the representativeness of F factors in the AAI equation described above and in the PA. 

These important uncertainties have been generally categorized as limitations in available field 

data as well as uncertainties that are related to reliance on the application of ecological and 

atmospheric modeling at the ecoregion scale to specify the various elements of the AAI.

With regard to data limitations, the Administrator observes that there are several 

important limitations in the available data upon which elements of the AAI are based. For 

example, while ambient measurements of NOy are made as part of a national monitoring 

network, the monitors are not located in locations that are representative of sensitive aquatic 

ecosystems. While air and water quality data are generally available in areas in the eastern U.S., 

there is relatively sparse coverage in mountainous western areas where a number of sensitive 

aquatic ecosystems are located. Further, even in areas where relevant data are available, small 

sample sizes impede efforts to characterize the representativeness of the available data, which 
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was noted by CASAC as being of particular concern. Also, measurements of reduced forms of 

nitrogen are available from only a small number of monitoring sites, and emission inventories for 

reduced forms of nitrogen used in atmospheric modeling are subject to considerable uncertainty.

With regard to uncertainties related to the use of ecological and atmospheric modeling, 

the Administrator notes in particular that model results are difficult to evaluate due to a lack of 

relevant observational data. For example, relatively large uncertainties are introduced by a lack 

of data with regard to pre-industrial environmental conditions and other parameters that are 

necessary inputs to critical load models that are the basis for factor F1 in the AAI equation. 

Also, observational data are not generally available to evaluate the modeled relationships 

between nitrogen and sulfur in the ambient air and associated deposition, which are the basis for 

the other factors (i.e., F2, F3, and F4) in the AAI equation.

In combination, these limitations and uncertainties result in a considerable degree of 

uncertainty as to how well the quantified elements of the AAI standard would predict the actual 

relationship between varying ambient concentrations of oxides of nitrogen and sulfur and steady

state ANC levels across the distribution of water bodies within the various ecoregions in the U.S. 

Because of this, there is considerable uncertainty as to the actual degree of protectiveness that 

such a standard would provide, especially for acid-sensitive ecoregions. The Administrator 

recognizes that the AAI equation, with factors quantified in the ranges discussed above and 

described more fully in the PA, generally performs well in identifying areas of the country that 

are sensitive to such acidifying deposition and indicates, as expected, that lower ambient levels 

of oxides of nitrogen and sulfur would lead to higher calculated AAI values. However, the 

uncertainties discussed here are critical for determining the actual degree of protection that 

would be afforded such areas by any specific target ANC level and percentile of water bodies 
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that would be chosen in setting a new AAI-based standard, and thus for determining an 

appropriate AAI-based standard that meets the requirements of section 109.

In considering these uncertainties, the Administrator notes that CASAC acknowledged 

that important uncertainties remain that would benefit from further study and data collection 

efforts, which might lead to potential revisions or modifications to the form of the standard 

developed in the PA. She also notes that CASAC encouraged the Agency to engage in future 

monitoring and model evaluation efforts to help inform the specification of model-derived 

elements in the AAI equation.

Based on the above considerations, the Administrator has determined that it is not 

appropriate under section 109 to set a new multi-pollutant standard to address deposition-related 

effects of oxides of nitrogen and sulfur on aquatic acidification at this time. Setting a NAAQS 

generally involves consideration of the degree of uncertainties in the science and other 

information, such as gaps in the relevant data and, in this case, limitations in the evaluation of the 

application of relevant ecological and atmospheric models at an ecoregion scale. As noted 

above, the issue here is not a question of uncertainties about the scientific soundness of the 

structure of the AAI, but instead uncertainties in the quantification and representativeness of the 

elements of the AAI as they vary in ecoregions across the country. At present, these 

uncertainties prevent an understanding of the degree of protectiveness that would be afforded to

various ecoregions across the country by a new standard defined in terms of a specific 

nationwide target ANC level and a specific percentile of water bodies for acid-sensitive 

ecoregions and thus prevent identification of an appropriate standard.. The Administrator has 

considered whether these uncertainties could be appropriately accounted for by choosing either a 

more or less protective target ANC level and percentile of water bodies than would otherwise be 
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chosen if the uncertainties did not substantially limit the confidence that can appropriately be 

ascribed to the quantification of the AAI elements. However, in the Administrator’s judgment, 

the uncertainties are of such nature and magnitude that there is no reasoned way to choose such a 

specific nationwide target ANC level or percentile of water bodies that would appropriately 

account for the uncertainties, since neither the direction nor the magnitude of change from the 

target level and percentile that would otherwise be chosen can reasonably be ascertained at this 

time. 

Based on the above considerations, the Administrator judges that the current limitations 

in relevant data and the uncertainties associated with specifying the elements of the AAI based 

on modeled factors are of such nature and degree as to prevent her from reaching a reasoned 

decision such that she is adequately confident as to what level and form (in terms of a selected 

percentile) of such a standard would provide any particular intended degree of protection of 

public welfare that the Administrator determined satisfied the requirements to set an appropriate 

standard under section 109. While acknowledging that CASAC supported moving forward to 

establish the standard developed in the PA, the Administrator also observes that CASAC 

supported conducting further field studies that would better inform the continued development or 

modification of such a standard. Given the large uncertainties and complexities inherent in 

quantifying the elements of such a standard, largely deriving from the unprecedented nature of 

the standard under consideration in this review, and having fully considered CASAC’s advice, 

the Administrator provisionally concludes that it is premature to set a new, multi-pollutant 

secondary standard for oxides of nitrogen and sulfur at this time, and as such she is proposing not 

to set such a new secondary standard.
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While it is premature to set such a multi-pollutant standard at this time, the Administrator 

determines that the Agency should undertake a field pilot program to gather additional data, and 

that it is appropriate that such a program be undertaken before, rather than after, reaching a 

decision to set such a standard. As described below in section IV, the purpose of the program is

to collect and analyze data so as to enhance our understanding of the degree of protectiveness 

that would likely be afforded by a standard based on the AAI as developed in the PA. This will 

provide additional information to aid the Agency in considering an appropriate multi-pollutant 

standard, specifically with respect to the acidifying effects of deposition of oxides of nitrogen 

and sulfur.PA. Data generated by this field program will also support development of an 

appropriate monitoring network that would work in concert with such a standard to result in the 

intended degree of protection. The data and analyses generated as a result of this program will 

serve to inform the next review of the NAAQS for oxides of nitrogen and sulfur. The 

information generated during the field program can also be used to help state agencies and EPA 

better understand how an AAI-based standard would work in terms of the implementation of 

such a standard.

Based on the above considerations, the Administrator is proposing not to set a new multipollutant AAI-based secondary standard for oxides of nitrogen and sulfur in this review. In 

reaching this decision, the Administrator recognizes that the new NO2 and SO2 primary 1-hour 

standards set in 2010, while not ecologically relevant for a secondary standard, will nonetheless 

result in reductions in oxides of nitrogen and sulfur that will directionally benefit the 

environment by reducing NOy and SOx deposition to sensitive ecosystems. EPA is proposing to 

revise the secondary standards by adding secondary standards identical to the NO2 and SO2

primary 1-hour standards set in 2010. More specifically, EPA is proposing a 1-hour secondary
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NO2 standard set at a level of 100 ppb and a 1-hour secondary SO2 standard set at a level of 75 

ppb. While this will not add secondary standards of an ecologically relevant form to address 

deposition-related effects, it will directionally provide some degree of additional protection. 

This is consistent with the view that the current secondary standards are neither sufficiently 

protective nor appropriate in form, but that it is not appropriate to propose to set a new, 

ecologically relevant multi-pollutant secondary standard at this time, for all of the reasons 

discussed above.

While not a basis for this decision, the Administrator also recognizes that a new, 

innovative AAI-based standard would raise significant implementation issues that would need to 

be addressed consistent with the CAA requirements for implementation-related actions following 

the setting of a new NAAQS. It will take time to address these issues, during which the Agency

will be conducting a field pilot program to gather relevant data and the environment will benefit 

from reductions in oxides of nitrogen and sulfur resulting from the new NO2 and SO2 primary

standards, as noted above, as well as reductions expected to be achieved from EPA’s Cross-State 

Air Pollution Rule and Mercury and Air Toxics standards. These implementation-related issues 

are discussed in more detail below in section IV.A.5.

The Administrator solicits comment on all aspects of this proposed decision, including 

the framework and elements of a multi-pollutant standard for oxides of nitrogen and sulfur to 

address deposition-related effects on sensitive ecosystems, with a focus on aquatic acidification, 

and the uncertainties and complexities associated with the development of such a standard at this 

time. The Administrator also solicits comment on the field pilot program and related monitoring 

methods as discussed below in section IV.

IV. Field Pilot Program and Ambient Monitoring 
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This section describes EPA’s plans for a field pilot program and the evaluation of 

monitoring methods for ambient air indicators of NOy and SOx to implement the Administrator’s 

decision to undertake such a field monitoring program in conjunction with her decision to 

propose not to set a new multi-pollutant secondary standard in this review, as discussed above in 

section III.H. As noted above and discussed below in section IV.A, the field pilot program is 

intended to collect and analyze data so as to enhance our understanding of the degree of 

protectiveness that would likely be afforded by a standard based on the AAI as developed in the PA. 

Data generated by this field program would also support development of an appropriate monitoring 

network that would work in concert with such a standard to result in the intended degree of 

protection. As discussed below in section IV.B, the evaluation of monitoring methods focuses on the 

development of Federal Reference Methods/Federal Equivalent Methods (FRM/FEM) for NOy

and SOx. The EPA notes that the monitoring program described here is intended to be 

coordinated with EPA’s CASTNET as a supplement to existing monitoring programs and is 

beyond the scope of the current CASTNET program. 

A. Field Pilot Program

This section presents the objectives of a field pilot program (section IV.A.1) that would 

gather relevant field data over a 5-year period in a sample of three to five sensitive ecoregions 

across the country. An overview of the scope and structure of the field program, with a focus on 

measurements of ambient air indicators of oxides of nitrogen and sulfur, is presented in section 

IV.A.2. Section IV.A.3 explains the role of additional complementary measurements beyond the 

ambient air indicators that would be included in the program, and section IV.A.4 discusses a 

parallel longer-term research agenda, both of which are guided by the uncertainties discussed

above in section III. Section IV.A.5 identifies implementation challenges presented by an AAI-
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based standard that could be addressed in parallel with a field pilot program. Section IV.A.6 

discusses engagement with stakeholder groups as part of the planned pilot program.

1. Objectives

Consideration of a new multi-pollutant standard to address deposition-related effects on 

sensitive aquatic ecoregions raises unique challenges relative to those typically raised in reviews 

of existing NAAQS for which an established network of FRM/FEM monitors, designed to 

measure the indicator pollutant, is generally available. The primary goal of this field pilot 

program, and the related monitoring program discussed in section IV.B, is to enhance our 

understanding of the degree of protectiveness that would likely be afforded by a standard based 

on the AAI, as described above in section III, so as to aid the Agency in considering an 

appropriate multi-pollutant standard that would be requisite to protect public welfare consistent 

with section 109 of the CAA, through the following objectives:

(1) Evaluate measurement methods for the ambient air indicators of NOy and SOx

and consider designation of such methods as FRMs;

(2) Examine the variability and improve characterization of concentration and 

deposition patterns of NOy and SOx, as well as reduced forms of nitrogen, within and 

across a number of sensitive ecoregions across the country;

(3) Develop updated ecoregion-specific factors (i.e., F1 through F4) for the AAI 

equation based in part on new observed air quality data within the sample ecoregions 

as well as on updated nationwide air quality model results and expanded critical load 

data bases, and explore alternative approaches for developing such representative 

factors;
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(4) Calculate ecoregion-specific AAI values using observed NOy and SOx data and 

updated ecoregion-specific factors to examine the extent to which the sample 

ecoregions would meet a set of alternative AAI-based standards;

(5) Develop air monitoring network design criteria for an AAI-based standard;

(6) assess the use of total nitrate measurements as a potential alternative indicator for

NOy;

(7) Support related longer-term research efforts, including enhancements to and 

evaluation of modeled dry deposition algorithms; and

(8) Facilitate stakeholder engagement in addressing implementation issues associated 

with possible future adoption of an AAI-based standard.

2. Overview of Field Pilot Program

The CASTNET program (Figure IV-1) affords an available infrastructure relevant to an 

AAI-based standard, given the location of sites in some acid-sensitive ecoregions and various 

measurements of sulfur and nitrogen species. The EPA plans to use CASTNET sites in selected 

acid-sensitive ecoregions to serve as the platform for this pilot program, potentially starting in 

late 2012 and extending through 2018. The CASTNET sites in three to five ecoregions in acidsensitive areas would collect NOy and SOx (i.e., SO2 and p-SO4) measurements over a 5-year 

period. The initial step in developing a data base of observed ambient air indicators for oxides of 

nitrogen and sulfur requires the addition of NOy samplers at the pilot study sites so that a full 

complement of indicator measurements are available to calculate AAI values. These CASTNET 

sites would also be used to make supplemental observations useful for evaluation of CMAQ’s 

characterization of factors F2 –F4 in the AAI equation.
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The selected ecoregions would account for geographic variability by including regions 

from across the U.S., including the east, upper midwest and west. Each selected region would 

have at least two existing CASTNET sites. Each of the pilot CASTNET sites would be used to 

evaluate the performance of the established methods, data retrieval and reporting procedures

used in the AAI equation.

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

 

Over the course of this 5-year pilot program, the most current national air quality

modeling, based on the most current national emissions inventory, would be used to develop an 
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updated set of F2 – F4 factors. A parallel multi-agency national critical load data base

development effort would be used as the basis for calculating updated F1 factors. As discussed 

above in section III.B, these factors would be based on average parameter values across an 

ecoregion. Using this new set of F factors, observations of NOy and SOx derived from the pilot 

program, averaged across each ecoregion, would be used to calculate AAI values in the sample 

ecoregions. The data from the pilot program would also be used to examine alternative 

approaches to generating representative air quality values, such as examining the appropriateness 

of spatial averaging in areas of high spatial variability.

3. Complementary Measurements

Complementary measurements may be performed at some sites in the pilot network to 

reduce uncertainties in the recommended methods and better characterize model performance 

and application to the AAI. The CASAC Air Monitoring and Methods Subcommittee (AMMS) 

advised EPA that such supplemental measurements were of critical importance in a field 

measurement program related to an AAI-based standard (Russell and Samet, 2011b).

Candidate complementary measurements to address sulfur, in addition to those provided 

by the CASTNET filter pack (CFP), include trace gas continuous SO2 and speciated PM2.5 

measurements. The co-located deployment of a continuous SO2 analyzer with the CFP for SO2

will provide test data for determining suitability of continuous SO2 measurements as a Federal 

Equivalent Method (FEM), as well as producing valuable time series data for model evaluation 

purposes. The weekly averaging time provided by the CFP adequately addresses the annualaverage basis of an AAI-based secondary standard, but would not be applicable to short-term 

(i.e., 1-hour) averages associated with the primary SO2 standard. Conversely, because of the low 

concentrations associated with many acid-sensitive ecoregions, existing SO2 Federal Reference 
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Methods (FRMs) designated for use in determining compliance with the primary standard would

not necessarily be appropriate for use in conjunction with an AAI-based secondary standard. 

Co-locating the PM2.5 sampler used in the EPA Chemical Speciation Network and the 

Interagency Monitoring of Protected Visual Environments (IMPROVE) network at pilot network 

sites would allow for characterizing the relationship between the CFP-derived p-SO4 and the 

speciation samplers used throughout the state and local air quality networks. Note that 

CASTNET already has several co-located IMPROVE chemical speciation samplers. Because 

the AAI equation is based on concentration of p-SO4, the original motivation for capturing all 

particle size fractions is not as important relative to simply capturing the concentration of total pSO4. 

Candidate measurements to complement oxidized nitrogen measurements, in addition to 

the CFP, include a mix of continuous and periodic sampling for the dominant NOy species, 

namely NO, true NO2, PAN, HNO3, and p-NO3. While there are several approaches to acquiring 

these measurements, perhaps the most efficient strategy would take advantage of the available 

CFP for total nitrate, and add a three-channel chemiluminescence instrument that will cycle 

between NOy, true NO2 and NO by adding photolytic detection for true NO2. Other options for 

measuring true NO2 would include adding either a stand-alone photolytic or cavity ring-down 

spectroscopy instrument. Measurements of PAN may be acquired either on a periodic basis 

through canister sampling and subsequent laboratory analysis or through emerging in-situ 

sampling and analysis methods. Although the CFP yields a reliable measurement of total nitrate, 

the t-NO3 (i.e., the sum of HNO3 and p-NH4) value, strong consideration may be given to direct 

measurement of HNO3, which has the highest deposition velocity of all the dominant NOy

species. Similar to the use of continuous SO2 data, these speciated NOy data serve two purposes:
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evaluating total NOy instrument behavior and evaluating air quality models. The measurement 

of individual NOy species can be used to generate site-specific NOy values for comparison to 

modeled NOy, and will likely provide insight into and improvement of modeled dry deposition.

The CASAC AMMS (Russell and Samet, 2011b) recommended that EPA consider the 

use of t-NO3 obtained from CASTNET sampling as an indicator for NOy, reasoning that t-NO3 is 

typically a significant fraction of deposited oxidized nitrogen in rural environments and 

CASTNET measurements are widely available. Collection of this data would support further

consideration of using the CFP for t-NO3 as the indicator of oxides of nitrogen for use in an AAIbased secondary standard. 

The CASAC AMMS also recommended that total NHx (NH3 and p-NH4) be considered 

as a proxy for reduced nitrogen species, reasoning that the subsequent partitioning to NH3 and pNH4 may be estimated using equilibrium chemistry calculations. Reduced nitrogen 

measurements are used to evaluate air quality modeling which is used in generating factor F2. 

Additional studies are needed to determine the applicability of NHx measurements and calculated 

values of NH3 and NH4 to the AAI. 

The additional supplemental measurements of speciated NOy, continuous SO2 and NHx

will be used in future air quality modeling evaluation efforts. Because there often is significant 

lag in the availability of contemporary emissions data to drive air quality modeling, the complete 

use of these data sets will extend beyond the 5-year collection period of the pilot program. 

Consequently, the immediate application of those data will address instrument performance 

comparisons that explore the feasibility of using continuous SO2 instruments in rural 

environments, and using the speciated NOy data to assess NOy instrument performance. 

Although contemporary air quality modeling will lag behind measurement data availability, the 
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observations can be used in deposition models to compare observed transference ratios with the 

previously calculated transference ratios to test temporal stability of the ratios. 

An extended water quality sampling effort should parallel the air quality measurement 

program to address some of the uncertainties related to factor F1 and the representativeness of 

the nth percentile critical load as discussed in section III.B.5.b.i. The objective of the water 

quality sampling would be to develop a larger data base of critical loads in each of the pilot 

ecoregions such that the nth percentile can adequately be characterized in terms of representing 

all water bodies. Opportunities to leverage and perhaps enhance existing ecosystem modeling 

efforts enabling more advanced critical load modeling and improved methods to estimate base 

cation production would be pursued. For example, areas with ongoing research studies 

producing data for dynamic critical load modeling would be considered when selecting the pilot 

ecoregions.

4. Complementary areas of research 

The EPA recognizes that a source of uncertainty in an AAI-based secondary standard that 

would not be directly addressed in the pilot program stems from the uncertainty in the model 

used to link atmospheric concentrations to dry deposition fluxes. Currently, there are no ongoing

direct dry deposition measurement studies at CASTNET sites that can be used to evaluate 

modeled results. It was strongly recommended by CASAC AMMS that a comprehensive 

sampling-intensive study be conducted in at least one, preferably two sites in different 

ecoregions to assess characterization of dry deposition of sulfur and nitrogen. These sites would 

be the same as those for the complementary measurements described above, but they would

afford an opportunity to also complement dry deposition process research that benefits from the 

ambient air measurements collected in the pilot program. The concerns regarding uncertainties 
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underlying an AAI-based secondary standard suggest that research that includes dry deposition 

measurements and evaluation of dry deposition models should be a high priority.

Similar leveraging should be pursued with respect to ecosystem research activities. For 

example, studies that capture a suite of soil, vegetation, hydrological, and water quality 

properties that can help evaluate more advanced critical load models would complement the 

atmospheric-based pilot program. In concept, such studies could provide the infrastructure for 

true multi-pollutant, multi-media “super” sites assuming the planning, coordination, and resource 

facets can be aligned. While this discussion emphasizes the opportunity of leveraging ongoing 

research efforts, consideration could be given to explicitly including related research components 

directly in the pilot program.

5. Implementation challenges

The CAA requires that once a NAAQS is established, designation and implementation 

must move forward. With a standard as innovative as the AAI-based standard considered in this 

review, the Administrator believes that its success will be greatly improved if, while additional 

data are being collected to reduce the uncertainties discussed above, the implementing agencies 

and other stakeholders have an opportunity to discuss and thoroughly understand how such a 

standard would work. And since, as noted above, emissions reductions that are directionally

correct to reduce aquatic acidification will be occurring as a result of other CAA programs, the 

Administrator believes that this period of further discussion will not delay progress but will 

ensure that once implementation is triggered, agencies will be prepared to implement it 

successfully.

Consideration of an AAI-based secondary standard for oxides of nitrogen and sulfur 

would present significant implementation challenges because it involves multiple, regionally-
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dispersed pollutants and relatively complex compliance determinations based on regionally 

variable levels of NOy and SOx concentrations that would be necessary to achieve a national 

ANC target.. The anticipated implementation challenges fall into three main categories: 

monitoring and compliance determinations for area designations, pre-construction permit 

application analyses of individual source impacts, and State Implementation Plan (SIP) 

development. Several overarching implementation questions that we anticipate will be addressed 

in parallel with the field pilot program’s five-year data collection period include: 

(1) What are the appropriate monitoring network density and siting requirements to 

support a compliance system based on ecoregions? 

(2) Given the unique spatial nature of the secondary standard (e.g., ecoregions), what are 

the appropriate parameters for establishing nonattainment areas?

(3) How can new or modified major sources of oxides of nitrogen and oxides of sulfur 

emissions assess their ambient impacts on the standard and demonstrate that they are not 

causing or contributing to a violation of the NAAQS for preconstruction permitting? To 

what extent does the fact that a single source may be impacting multiple areas, with 

different acid sensitivities and variable levels of NOy and SOx concentrations that would 

be necessary to achieve a national ANC target, complicate this assessment and how can 

these additional complexities best be addressed?

(4) What additional tools, information, and planning structures are needed to assist states 

with SIP development, including the assessment of interstate pollutant transport and 

deposition?

(5) Would transportation conformity apply in nonattainment and maintenance areas for 

this secondary standard, and, if it does, would satisfying requirements that apply for 
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related primary standards (e.g., ozone, PM2.5, and NO2) be demonstrated to satisfy

requirements for this secondary standard?

6. Final monitoring plan development and stakeholder participation

The existing CASTNET sampling site infrastructure provides an effective means of 

quickly and efficiently deploying a monitoring program to support potential implementation of 

an AAI-based secondary standard, and also provides an additional opportunity for federally 

managed networks to collaborate and support the states, local agencies and tribes (SLT) in 

determining compliance with a secondary standard. A collaborative effort would help to 

optimize limited federal and SLT monitoring funds and would be beneficial to all involved. The 

CASTNET is already a stakeholder-based program with over 20 participants and contributors, 

including federal, state and tribal partners.

The CASAC AMMS generally endorsed the technical approaches used in CASTNET, but 

concerns were raised by individual representatives of state agencies concerning the perception of 

EPA-controlled management aspects of CASTNET and data ownership. Potential approaches to 

resolve these issues will be developed and evaluated in existing National Association of Clean

Air Agencies (NACAA)/EPA ambient air monitoring workgroups. The EPA Office of Air and 

Radiation (which includes the Office of Air Quality Planning Standards, OAQPS; and the Office 

of Atmospheric Program’s Clean Air Markets Division, OAP-CAMD), and their partners on the 

NACAA monitor steering committee will develop a prioritized specific plan that identifies the 

three to five ecoregions and the instrumentation to be deployed. The EPA anticipates that a cost 

estimate of the plan with priorities and options will be developed by January, 2012. Although 

this pilot program is focused on data collection, the plan will include details of the data analysis 
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approaches as well as a vehicle that incorporates engagement from those within EPA and SLTs 

to foster progress on the implementation questions noted above in section IV.A.5.

If an AAI-based secondary standard were to be set in the future, deployment of a full 

national network would follow the pilot monitoring program. The number of sites deployed in 

the network will lead to increased confidence in capturing spatial patterns of air quality. 

Recognizing that this section presents the general elements of the field pilot programs, EPA 

intends to develop a more detailed field pilot program plan through a process that will engage the 

air quality management and research (atmospheric and ecosystem) communities, as well as other 

federal agencies, state and local agencies, and non-government based centers of expertise. The 

EPA is seeking comment and input on all aspects of this field pilot program.

.B. Evaluation of Monitoring Methods

The EPA generally relies on monitoring methods that have been designated as FRMs or 

FEMs for the purpose of determining the attainment status of areas with regard to existing 

NAAQS. Such FRMs or FEMs are generally required to measure the air quality indicators that 

are compared to the level of a standard to assess compliance with a NAAQS. Prior to their 

designation by EPA as FRM/FEMs through a rulemaking process, these methods must be 

determined to be applicable for routine field use and need to have been experimentally validated 

by meeting or exceeding specific accuracy, reproducibility, and reliability criteria established by 

EPA for this purpose. As discussed above in section III.A, the ambient air indicators being 

considered for use in an AAI-based standard include SO2, particulate sulfate (p-SO4), and total 

reactive oxides of nitrogen (NOy).

The CASTNET provides a well established infrastructure that would meet the basic 

location and measurement requirements of an AAI-based secondary standard given the rural 
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placement of sites in acid sensitive areas. In addition, CFPs currently provide very economical 

weekly, integrated average concentration measurements of SO2, p-SO4, ammonium ion (NH4) 

and t-NO3, the sum of HNO3 and p-NO3. 

While routinely operated instruments that measure SO2, p-SO4, NOy and/or t-NO3 exist, 

instruments that measure p-SO4, NOy, t-NO3, or the CFP for SO2 have not been designated by 

EPA as FRMs or FEMs. The EPA’s Office of Research and Development has initiated work that 

will support future FRM designations by EPA for SO2 and p-SO4 measurements based on the 

CFP. Such a designation by EPA could be done for the purpose of facilitating consistent 

research related to an AAI-based standard and/or in conjunction with setting and supporting an 

AAI-based secondary standard. 

Based on extensive review of literature and available data, the EPA has identified 

potential methods that appear suitable for measuring each of the three components of the 

indicators. These three methods are being considered as new FRMs to be used for measuring the 

ambient concentrations of the three components that would be needed to determine compliance 

with an AAI-based secondary standard.

For the SO2 and p-SO4 measurements, EPA is considering the CFP method, which 

provides weekly average concentration measurements for SO2 and p-SO4. This method has been 

used in the EPA’s CASTNET monitoring network for 15 years, and strongly indicates that it will 

meet the requirements for use as an FRM for the SO2 and p-SO4 concentrations for an AAI-based 

secondary standard. 

Although the CFP method would provide measurements of both the SO2 and p-SO4 

components in a unified sampling and analysis procedure, individual FRMs will be considered 

for each. The EPA recognizes that an existing FRM to measure SO2 concentrations using ultra-
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violet fluorescence (UVF) exists (40 CFR Part 50, Appendix A-1) for the purpose of monitoring 

compliance for the primary SO2 NAAQS. However, several factors suggest that the CFP method 

would be superior to that UVF FRM for monitoring compliance with an AAI-based secondary

standard and will be discussed in more detail below. 

For monitoring the NOy component, a continuous analyzer for measuring NOy is 

commercially available and is considered to be suitable for use as an FRM. This method is 

similar in design to the existing NO2 FRM (described in 40 CFR Part 50, Appendix F), which is 

based on the ozone chemiluminescence measurement technique. The method is adapted to and 

further optimized to measure all NOy. However, this NOy method requires further evaluation 

before it can be fully confirmed as a suitable FRM. The EPA is currently completing a full 

scientific assessment of the NOy method to determine whether it would be appropriate to 

consider for designation by EPA as an FRM. Specific details on these three methods are given 

below. 

 On February 16, 2011, EPA presented this set of potential FRMs to the CASAC AMMS 

for their consideration and comment. In response, the CASAC AMMS stated that, overall, it 

believes that EPA’s planned evaluation of methods for measuring NOy, SO2 and p-SO4 as 

ambient air indicators is a suitable approach in concept. On supporting the CFP method as a 

potential FRM for SO2, CASAC stated that they felt that the CFP is adequate for measuring 

long-term average SO2 gas concentrations in rural areas with low levels (less than 5 parts per 

billion by volume (ppbv)) and is therefore suitable for consideration as an FRM. For p-SO4, 

CASAC generally supports the use of the CFP as a potential FRM for measuring p-SO4 for an 

AAI-based secondary standard. The method has been relatively well-characterized and 
Page 205 of 238

evaluated, and it has a documented, long-term track record of successful use in a field network 

designed to assess spatial patterns and long-term trends.

On supporting the photometric NOy method as a potential FRM, CASAC concluded that 

the existing NOy method is generally an appropriate approach for the indicator. However, 

CASAC agrees that additional characterization and research is needed to fully understand the 

method in order to designate it as a FRM. The EPA is now soliciting public comment on these 

methods as to their adequacy, suitability, and relative merits as FRMs for purposes of monitoring 

to determine compliance with an AAI-based secondary standard.

1. Potential FRMs for SO2 and p-SO4

The CFP is a combined, integrated sampling and analysis method based on the wellestablished measurement technology that has been used extensively in EPA’s CASTNET 

monitoring network (see http://www.epa.gov/castnet). This method is in current use at over 80 

monitoring sites and has been in use at not less than 40 sites for over 15 years. This method 

employs a relatively simple and inexpensive sampler and uses four 47-mm filters placed in an 

open-faced filter pack to simultaneously collect integrated filter samples for the SO2 and p-SO4

components. In addition, the CFP is also capable of the collection of t-NO3, the sum of HNO3

and p-NO3.

The first stage of the filter pack assembly contains a Teflon® filter that collects p-SO4

2-

and p-NO3, the second stage contains a nylon filter that collects SO2 (as SO4

2-

) and HNO3, and 

the third stage contains two cellulose fiber filters impregnated with potassium carbonate (K2CO3) 

that collect any remaining SO2 (as SO4

2-

). The sampler collects 1-week integrated samples at a 

very low, controlled flow rate (1.5 or 3 L/min) in an attempt simulate actual deposition. Weekly
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averaged SO2 and p-SO4 concentrations could then be averaged over a 1-year period to calculate 

annual average values.

Upon sample completion, the species-specific filters are extracted, with subsequent 

analysis by the well-established and documented ion chromatographic (IC) analytical technique. 

During the IC analysis, an aliquot of a filter extract is injected into a stream of eluent (ion 

chromatography mobile phase, generally a millimolar-strength solution of carbonatebicarbonate) and passed through a series of ion exchangers. The anions of interest are separated 

on the basis of their relative affinities for a low capacity and the strongly basic anion exchanger 

(guard and separator column). The separated anions are directed onto a cation exchanger 

(suppressor column) where they are converted to their highly conductive acid form, and the 

eluent is converted to a weakly conductive form. The now-separated anions, each in their acid 

form, are measured by conductivity. They are identified on the basis of retention time compared 

to that of standards and quantified by measurement of peak area compared to the peak areas of 

calibration standards. 

Calibration and quality assurance for the method are applied to the sample filters, the 

analytical processes, and the flow rate measurement and control aspects of the sampler. Overall 

method performance is typically assessed with collocated samplers. These quality assurance 

techniques are routinely used and have proved adequate for other types of FRMs and equivalent 

methods in air monitoring network service.

The measurement and analytical procedures and past performance data associated with 

the CFP method are well documented and available through Quality Assurance Performance 

Plans (QAPPs), Standard Operating Procedures (SOPs) and annual reports (US EPA, 2010a and 

2010b). The accumulated database on the CFP method is substantial and indicates that the 
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method is sound, stable and has good reliability in routine, field operation. Data quality 

assessment results show the method to have good reproducibility, with collocated and analytical 

precision values in the range of 2 percent to 10 percent (excluding very low concentration 

measurements near the method detection limits; US EPA 2010b).

Data quality objectives (DQOs) for a new FRM would be based upon current DQOs 

being used for this method by EPA’s OAP/CAMD and the NPS, the federal managers of 

CASTNET (US EPA, 2010a). In its current state, the CFP method is expected to meet or exceed 

(as past CASTNET data have indicated; US EPA, 2010b) the expected FRM DQOs, even when 

deployed in new monitoring networks outside of CASTNET. In addition, CASTNET samples 

have agreed favorably with other measures of SO2 and p-SO4 in comparison studies. For 

example, in direct comparison with an annular denuder sampler (ADS) method, CASTNET/ADS 

ratios for SO2 and p-SO4 were generally on the order of 0.9-1.1 (Lavery et al, 2009; Sickles et al, 

1999; Sickles et al, 2008), thus illustrating the accuracy of the CFP method in the determination 

of long-term average SO2 and p-SO4 concentrations. The EPA believes that the CFP method 

would be fully adequate as an FRM in determining yearly average SO2 and p-SO4 concentrations 

for compliance determination purposes.

The EPA recognizes that an existing FRM for SO2 has proven adequate for the purposes 

of monitoring compliance for the primary SO2 NAAQS, specifically the newly-promulgated 1-

hour standard. However, this FRM is better suited to the shorter-term, higher concentration 

primary and secondary SO2 NAAQS, and there is substantial uncertainty as to the adequacy of 

this SO2 FRM for monitoring the lower concentrations relevant to determining compliance with 

an AAI-based secondary standard. The performance specifications for SO2 FRM analyzers (40 

CFR Part 53, Table B-1) require a lower detectable limit (LDL) of 0.002 ppm for the standard 
Page 208 of 238

measurement range and 0.001 ppm for the lower measurement range. These requirements 

correspond to mass per unit volume concentrations of 5.24 and 2.62 μg/m3, respectively. 

Analysis of 2009 CASTNET data shows that of the 84 CASTNET sampling sites, 63 measured 

annual average SO2 concentrations below even the lower of these LDL requirements of 2.62 

μg/m3for the lower range SO2 FRM (US EPA, 2010a). In addition, 11 of the 84 sites measured 

annual (2009) average SO2 concentrations very near or below the manufacturers’ reported 

detection limits for trace level UVF SO2 monitors. Further, it is likely that the number of sites 

with annual average SO2 concentration below both the SO2 FRM LDL and the manufacturers 

reported detection limits will increase due to expected declines in mean SO2 concentrations (US

EPA, 2010b). For these reasons, EPA is considering the CFP method for use as the FRM for 

monitoring the SO2 component of an ambient air indicator for oxides of sulfur, with a 

recommendation for additional study and data collection to evaluate further the possible 

applicability of the continuous UVF SO2 FRM for this purpose.

2. Potential FRM for NOy

Atmospheric concentrations of NOy are measured continuously by an analyzer that 

photometrically measures the light intensity, at wavelengths greater than 600 nanometers (nm), 

resulting from the chemiluminescent reaction of ozone (O3) with NO in sampled air. This 

method is very similar to the chemiluminescence NO/NO2 analyzers widely used to collect NO2

monitoring data for determining compliance with the NO2 NAAQS. The various oxides of 

nitrogen species, excluding NO, are first quantitatively reduced to NO by means of a catalytic 

converter. These species include NO2, HNO2, PANs, HNO3 and p-NO3. The NO, which 

commonly exists in ambient air, passes through the converter unchanged, and, when combined 

with the NO resulting from the catalytic conversion of the other oxides of nitrogen, a 
Page 209 of 238

measurement of the total NOy concentration results. To maximize the conversion of the more 

chemically active oxides of nitrogen species, the converter is located externally, at or near the air 

sample inlet probe. This location minimizes losses of these active species that could otherwise 

occur from chemical reactions and wall losses in the sample inlet line.

The NOy analyzer is a suitable, commercially produced continuous chemiluminescence 

analyzer that includes an ozone generator, a reaction cell, a photometric detector, wavelength 

filters as necessary to reduce sensitivity to wavelengths below 600 nanometer (nm), a pump and 

flow control system to draw atmospheric air through the converter and into the reaction cell, a 

suitable converter, a system to control the operation of the analyzer, and appropriate electronics 

to process and quantitatively scale the photometric signals. The converter contains a catalyst 

such as molybdenum and is heated to an optimum temperature designed to optimize the 

conversion of the various oxides of nitrogen to NO. It is connected to the analyzer via suitable 

lengths of Teflon® tubing. Hourly NOy measurements obtained by the analyzer would be 

averaged over the same 7-day period used by the CFP method to measure the SO2 and p-SO4

components, with further averaging over a 1-year period.

Commercial NOy analyzers are currently available, and the analyzers have been used for 

a variety of monitoring applications. During the 2006 TexAQS Radical and Aerosol 

Measurement Project (TRAMP), Luke et al., 2010, compared measured NOy concentrations 

obtained with an NOy instrument based upon the above mentioned methodology with the sum of 

measured individual NOy species (i.e., NOyi = NO+NO2+HNO3+PANs+HNO2+p-NO3). This 

comparison yielded excellent overall agreement during both day ([NOy](ppb) = [NOyi](ppb) × 

1.03 – 0.42; r2= 0.9933) and night time ([NOy](ppb) = [NOyi](ppb) × 1.01 – 0.18; r2= 0.9975) 

periods (Luke et al, 2010). The results of this study show that this NOy method is capable of the 
Page 210 of 238

accurate determination of all the atmospherically relevant NOy components, resulting in an 

accurate determination of total NOy concentrations. The NOy instruments have been routinely 

operated in networks such as SouthEastern Aerosol Research and Characterization (SEARCH), 

dating back several years. In addition, state monitoring agencies across the U.S. have begun, 

starting in 2009, the routine operation of commercially available NOy instrumentation in 

anticipation of EPA’s NCore network transitioning to full operation in 2011. 

These initial assessments described above are promising and indicate that the photometric 

NOy method appears to be accurate, reliable, and capable of routine network operation. As a 

result, the method is most likely capable for use as an FRM for determining atmospheric NOy

concentrations as a component in determining compliance with an AAI-based secondary

standard. However, as described below, this continuous method for NOy requires additional time 

for further evaluation before it can be fully confirmed for adoption as a FRM. The EPA has 

identified measurement uncertainties and some remaining science questions associated with this 

method. Among these are: (a) the ability of the method to capture all components of NOy

relevant to nitrogen deposition, (b) the efficiency of the molybdenum converter in converting all 

oxides of nitrogen to NO for detection (excluding NO2, as this conversion is already well 

documented), (c) appropriate inlet height specifications to minimize any bias associated with 

vertical concentration gradients of key NOy components, (d) identification and quantification of 

potential measurement interferences in the NOy determination, and (e) development and 

demonstration of effective calibration/challenge procedures to best represent the various 

mixtures of NOy components that are expected to be present in the different air sheds across the 

U.S.
Page 211 of 238

To address these NOy method uncertainties and to fully assess this method for use as the 

NOy FRM, EPA has developed a detailed research plan (Russell and Samet, 2011b) which was 

presented to the CASAC AMMS on February 16, 2011. In response, CASAC recognized the 

need for, and supported the general outline of EPA’s research plan to evaluate the NOy method 

for potential designation as an FRM (US EPA, 2011). In addition, the CASAC AMMS 

suggested additional areas of research associated with the photometric NOy method that warrant 

further assessment prior to final designation of the method as the NOy FRM. These include 

operation of the method during extremely low temperature conditions to investigate possible 

condensation in sample lines, method detection limits relative to low levels expected in remote 

areas, and ambient-based method evaluations in various air sheds across the U.S. In response to 

these CASAC AMMS suggestions, EPA is carrying out studies, in addition to the tasks outlined 

in the research plan, for the NOy method. The results of these studies will likely take a year or 

more to become available. As noted previously, EPA anticipates that these results will be 

favorable and will confirm the adequacy of the NOy method as a suitable FRM for determining 

compliance with an AAI-based secondary standard.

V. Statutory and Executive Order Reviews

A. Executive Order 12866: Regulatory Planning and Review and Review and Executive 

Order 13563: Improving Regulation and Regulatory Review

Under Executive Order 12866 (58 FR 51735, October 4, 1993), this action is a

“significant regulatory action.” Accordingly, EPA submitted this action to the Office of 

Management and Budget (OMB) for review under Executive Orders 12866 and 13563 (76 FR 

3821, January 21, 2011), and any changes made in response to OMB recommendations have 

been documented in the docket for this action.
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B. Paperwork Reduction Act

This action does not impose an information collection burden under the provisions of the 

Paperwork Reduction Act, 44 U.S.C. 3501 et seq. Burden is defined at 5 CFR 1320.3(b). There 

are no information collection requirements directly associated with the establishment of a 

NAAQS under section 109 of the CAA.

C. Regulatory Flexibility Act

For purposes of assessing the impacts of today’s rule on small entities, small entity is 

defined as: (1) a small business that is a small industrial entity as defined by the Small Business 

Administration’s (SBA) regulations at 13 CFR 121.201; (2) a small governmental jurisdiction 

that is a government of a city, county, town, school district or special district with a population of 

less than 50,000; and (3) a small organization that is any not-for-profit enterprise which is 

independently owned and operated and is not dominant in its field.

After considering the economic impacts of today's proposed rule on small entities, I

certify that this action will not have a significant economic impact on a substantial number of 

small entities. This proposed rule will not impose any requirements on small entities. Rather, 

this rule establishes national standards for allowable concentrations of oxides of nitrogen and 

sulfur in ambient air as required by section 109 of the CAA. See also American Trucking 

Associations v. EPA. 175 F. 3d at 1044-45 (NAAQS do not have significant impacts upon small 

entities because NAAQS themselves impose no regulations upon small entities). We continue to 

be interested in the potential impacts of the proposed rule on small entities and welcome 

comments on issues related to such impacts.

D. Unfunded Mandates Reform Act
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Title II of the Unfunded Mandates Reform Act of 1995 (UMRA), Public Law 104-4, 

establishes requirements for Federal agencies to assess the effects of their regulatory actions on 

State, local, and Tribal governments and the private sector. Under section 202 of the UMRA, 

EPA generally must prepare a written statement, including a cost-benefit analysis, for proposed 

and final rules with “Federal mandates” that may result in expenditures to state, local, and tribal 

governments, in the aggregate, or to the private sector, of $100 million or more in any 1 year. 

Before promulgating an EPA rule for which a written statement is needed, section 205 of the 

UMRA generally requires EPA to identify and consider a reasonable number of regulatory

alternatives and to adopt the least costly, most cost-effective or least burdensome alternative that 

achieves the objectives of the rule. The provisions of section 205 do not apply when they are 

inconsistent with applicable law. Moreover, section 205 allows EPA to adopt an alternative 

other than the least costly, most cost-effective or least burdensome alternative if the 

Administrator publishes with the final rule an explanation why that alternative was not adopted. 

Before EPA establishes any regulatory requirements that may significantly or uniquely affect 

small governments, including tribal governments, it must have developed under section 203 of 

the UMRA a small government agency plan. The plan must provide for notifying potentially 

affected small governments, enabling officials of affected small governments to have meaningful 

and timely input in the development of EPA regulatory proposals with significant Federal 

intergovernmental mandates, and informing, educating, and advising small governments on 

compliance with the regulatory requirements. 

This action contains no Federal mandates under the provisions of Title II of the Unfunded 

Mandates Reform Act of 1995 (UMRA), 2 U.S.C. 1531-1538 for state, local, or tribal 

governments or the private sector. Therefore, this action is not subject to the requirements of 
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sections 202 or 205. Furthermore, as indicated previously, in setting a NAAQS EPA cannot 

consider the economic or technological feasibility of attaining ambient air quality standards; 

although such factors may be considered to a degree in the development of state plans to 

implement the standards. See also American Trucking Associations v. EPA, 175 F. 3d at 1043 

(noting that because EPA is precluded from considering costs of implementation in establishing 

NAAQS, preparation of a Regulatory Impact Analysis pursuant to the Unfunded Mandates 

Reform Act would not furnish any information which the court could consider in reviewing the 

NAAQS). Accordingly, EPA has determined that the provisions of sections 202, 203, and 205 of 

the UMRA do not apply to this proposed decision. The EPA acknowledges, however, that any

corresponding revisions to associated state implementation plan (SIP) requirements and air 

quality surveillance requirements, 40 CFR part 51 and 40 CFR part 58, respectively, might result 

in such effects. Accordingly, EPA will address, as appropriate, unfunded mandates if and when 

it proposes any revisions to 40 CFR parts 51 or 58.

E. Executive Order 13132: Federalism

This proposed rule does not have federalism implications. It will not have substantial 

direct effects on the states, on the relationship between the national government and the states, or 

on the distribution of power and responsibilities among the various levels of government, as 

specified in Executive Order 13132 because it does not contain legally binding requirements. 

Thus, the requirements of Executive Order 13132 do not apply to this rule.

EPA believes, however, that this proposed rule may be of significant interest to state 

governments. As also noted in section E (above) on UMRA, EPA recognizes that states will 

have a substantial interest in this rule and any corresponding revisions to associated SIP 

requirements and air quality surveillance requirements, 40 CFR part 51 and 40 CFR part 58, 
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respectively. Therefore, in the spirit of Executive Order 13132 and consistent with EPA policy 

to promote communications between EPA and state and local governments, EPA specifically 

solicits comment on this proposed rule from state and local officials.

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

Executive Order 13175, entitled “Consultation and Coordination with Indian Tribal 

Governments” (65 FR 67249, November 9, 2000), requires EPA to develop an accountable 

process to ensure “meaningful and timely input by tribal officials in the development of 

regulatory policies that have tribal implications.” This rule concerns the establishment of 

national standards to address the public welfare effects of oxides of nitrogen and sulfur.

This action does not have Tribal implications, as specified in Executive Order 13175 (65 

FR 67249, November 9, 2000). It does not have a substantial direct effect on one or more Indian 

tribes, since tribes are not obligated to adopt or implement any NAAQS. Thus, Executive Order 

13175 does not apply to this rule.

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

Risks

This action is not subject to EO 13045 because it is not an economically significant rule 

as defined in EO 12866. 

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

Use

This action is not a “significant energy action” as defined in Executive Order 13211 (66 

FR 28355, May 22, 2001), because it is not likely to have a significant adverse effect on the 

supply, distribution, or use of energy. This action concerns the establishment of national 

standards to address the public welfare effects of oxides of nitrogen and sulfur. This action does 
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not prescribe specific pollution control strategies by which these ambient standards will be met. 

Such strategies will be developed by states on a case-by-case basis, and EPA cannot predict 

whether the control options selected by states will include regulations on energy suppliers, 

distributors, or users.

I. National Technology Transfer and Advancement Act

Section 12(d) of the National Technology Transfer and Advancement Act of 1995 

(NTTAA), Public Law No. 104-113, 12(d) (15 U.S.C. 272 note) directs EPA to use voluntary

consensus standards in its regulatory activities unless to do so would be inconsistent with 

applicable law or otherwise impractical. Voluntary consensus standards are technical standards 

(e.g., materials specifications, test methods, sampling procedures, and business practices) that are 

developed or adopted by voluntary consensus standards bodies. The NTTAA directs EPA to

provide Congress, through OMB, explanations when the Agency decides not to use available and 

applicable voluntary consensus standards.

The EPA is not aware of any voluntary consensus standards that are relevant to the 

provisions of this proposed rule. The EPA welcomes any feedback on such standards that may

be applicable.

J. Executive Order 12898: Federal Actions to Address Environmental Justice in Minority 

Populations and Low-Income Populations

Executive Order 12898 (59 FR 7629 (Feb. 16, 1994)) establishes federal executive policy

on environmental justice. Its main provision directs federal agencies, to the greatest extent 

practicable and permitted by law, to make environmental justice part of their mission by

identifying and addressing, as appropriate, disproportionately high and adverse human health or 
Page 217 of 238

environmental effects of their programs, policies, and activities on minority populations and lowincome populations in the United States. 

EPA has determined that this proposed rule will not have disproportionately high and 

adverse human health or environmental effects on minority or low-income populations because it 

retains the level of environmental protection for all affected populations without having any 

disproportionately high and adverse human health or environmental effects on any population, 

including any minority or low-income population.

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Page 220 of 238 – NOx/SOx Secondary NAAQS

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List of Subjects in 40 CFR Part 50

Environmental protection, Air pollution control, Carbon monoxide, Lead, Nitrogen 

dioxide, Ozone, Particulater matter, Sulfur oxides.

________________________________

Dated: July 12, 2011

Lisa P. Jackson,

Administrator
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For the reasons set forth in the preamble, part 50 of chapter 1 of title 40 of the code of Federal 

regulations is proposed to be amended as follows:

PART 50-NATIONAL PRIMARY AND SECONDARY AMBIENT AIR QUALITY 

STANDARDS 

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

Authority: 42 U.S.C. 7401, et seq.

2. Section 50.5 is amended by revising paragraphs (b) and (c) and by adding paragraphs (d) and 

(e) to read as follows:

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

* * * * *

(b) The level of the national secondary 1-hour ambient air quality standard for oxides of sulfur is 

75 parts per billion (ppb, which is 1 part in 1,000,000,000), measured in the ambient air as sulfur 

dioxide (SO2).

(c) The levels of the standards shall be measured by a reference method based on Appendix A-1 

or A-2 of this part, or by a Federal Equivalent Method (FEM) designated in accordance with part 

53 of this chapter.

(d) To demonstrate attainment with the 3-hour secondary standard, the second-highest 3-hour 

average must be based upon hourly data that are at least 75 percent complete in each calendar 

quarter. A 3-hour block average shall be considered valid only if all three hourly averages for 
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the 3-hour period are available. If only one or two hourly averages are available, but the 3-hour 

average would exceed the level of the standard when zeros are substituted for the missing values, 

subject to the rounding rule of paragraph (a) of this section, then this shall be considered a valid 

3-hour average. In all cases, the 3-hour block average shall be computed as the sum of the 

hourly averages divided by 3.

(e) The 1-hour secondary standard is met at an ambient air quality monitoring site when the 

three-year average of the annual 99th percentile of the daily maximum 1-hour average 

concentrations is less than or equal to 75 ppb, as determined in accordance with Appendix T of 

this part.

3. Section 50.11 is amended by deleting paragraphs (f) and (g) and revising paragraphs (a) 

through (e) to read as follows: 

§50.11 National primary and secondary ambient air quality standards for oxides of 

nitrogen (with nitrogen dioxide as the indicator).

 (a) The level of the national primary and secondary annual ambient air quality standards for 

oxides of nitrogen is 53 parts per billion (ppb, which is 1 part in 1,000,000,000), annual average 

concentration, measured in the ambient air as nitrogen dioxide.

(b) The level of the national primary and secondary 1-hour ambient air quality standards for 

oxides of nitrogen is 100 ppb, 1-hour average concentration, measured in the ambient air as 

nitrogen dioxide. 

(c) The levels of the standards shall be measured by:
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(1) a reference method based on appendix F to this part; or 

(2) a Federal equivalent method (FEM) designated in accordance with part 53 of this chapter.

(d) The annual primary and secondary standards are met when the annual average concentration 

in a calendar year is less than or equal to 53 ppb, as determined in accordance with Appendix S

of this part for the annual standard.

(e) The 1-hour primary and secondary standards are met when the three-year average of the 

annual 98th percentile of the daily maximum 1-hour average concentration is less than or equal to 

100 ppb, as determined in accordance with Appendix S of this part for the 1-hour standard.

4. Appendix S is amended as follows:

a. by revising paragraph 1.(a),

b. by revising the definition of “Design values” under paragraph 1.(c),

c. by revising paragraph 2.(b),

d. by revising paragraphs 3.1(a) through (d),

e. by revising paragraphs 3.2(a) through (e),

f. by revising paragraph 4.1(b),

g. by revising paragraph 4.2 (c),

h. by revising paragraph 5.1(b), and
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i. by revising paragraph 5.2(b) to read as follows:

Appendix S to Part 50—Interpretation of the Primary and Secondary National Ambient 

Air Quality Standards for Oxides of Nitrogen (Nitrogen Dioxide) 

1. General.

(a) This appendix explains the data handling conventions and computations necessary for 

determining when the primary and secondary national ambient air quality standards for oxides of 

nitrogen as measured by nitrogen dioxide (“NO2 NAAQS”) specified in §50.11 are met. 

Nitrogen dioxide (NO2) is measured in the ambient air by a Federal reference method (FRM) 

based on appendix F to this part or by a Federal equivalent method (FEM) designated in 

accordance with part 53 of this chapter. Data handling and computation procedures to be used in 

making comparisons between reported NO2 concentrations and the levels of the NO2 NAAQS 

are specified in the following sections. 

* * * * * 

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

* * * * * 

Design values are the metrics (i.e., statistics) that are compared to the NAAQS levels to 

determine compliance, calculated as specified in section 5 of this appendix. The design values 

for the primary and secondary NAAQS are:

(1) The annual mean value for a monitoring site for one year (referred to as the “annual primary

or secondary standard design value”). 
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(2) The 3-year average of annual 98th percentile daily maximum 1-hour values for a monitoring 

site (referred to as the “1-hour primary or secondary standard design value”).

* * * * * 

2. Requirements for Data Used for Comparisons with the NO2 NAAQS and Data

Reporting Considerations.

* * * * * 

(b) When two or more NO2 monitors are operated at a site, the state may in advance designate 

one of them as the primary monitor. If the state has not made this designation, the Administrator 

will make the designation, either in advance or retrospectively. Design values will be developed 

using only the data from the primary monitor, if this results in a valid design value. If data from 

the primary monitor do not allow the development of a valid design value, data solely from the 

other monitor(s) will be used in turn to develop a valid design value, if this results in a valid 

design value. If there are three or more monitors, the order for such comparison of the other 

monitors will be determined by the Administrator. The Administrator may combine data from 

different monitors in different years for the purpose of developing a valid 1-hour primary or 

secondary standard design value, if a valid design value cannot be developed solely with the data 

from a single monitor. However, data from two or more monitors in the same year at the same 

site will not be combined in an attempt to meet data completeness requirements, except if one 

monitor has physically replaced another instrument permanently, in which case the two 

instruments will be considered to be the same monitor, or if the state has switched the 

designation of the primary monitor from one instrument to another during the year.
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* * * * * 

3. Comparisons with the NO2 NAAQS.

3.1 The Annual Primary and Secondary NO2 NAAQS.

(a) The annual primary and secondary NO2 NAAQS are met at a site when the valid annual 

primary standard design value is less than or equal to 53 parts per billion (ppb).

(b) An annual primary or secondary standard design value is valid when at least 75 percent of the 

hours in the year are reported. 

(c) An annual primary or secondary standard design value based on data that do not meet the 

completeness criteria stated in section 3.1(b) may also be considered valid with the approval of, 

or at the initiative of, the Administrator, who may consider factors such as monitoring site 

closures/moves, monitoring diligence, the consistency and levels of the valid concentration 

measurements that are available, and nearby concentrations in determining whether to use such 

data. 

(d) The procedures for calculating the annual primary and secondary standard design values are 

given in section 5.1 of this appendix.

3.2 The 1-hour Primary and Secondary NO2 NAAQS.

(a) The 1-hour primary or secondary NO2 NAAQS is met at a site when the valid 1-hour primary

or secondary standard design value is less than or equal to 100 parts per billion (ppb).

(b) An NO2 1-hour primary or secondary standard design value is valid if it encompasses three 

consecutive calendar years of complete data. A year meets data completeness requirements 

when all 4 quarters are complete. A quarter is complete when at least 75 percent of the sampling 

days for each quarter have complete data. A sampling day has complete data if 75 percent of the 
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hourly concentration values, including state-flagged data affected by exceptional events which 

have been approved for exclusion by the Administrator, are reported.

(c) In the case of one, two, or three years that do not meet the completeness requirements of 

section 3.2(b) of this appendix and thus would normally not be useable for the calculation of a 

valid 3-year 1-hour primary or secondary standard design value, the 3-year 1-hour primary or 

secondary standard design value shall nevertheless be considered valid if one of the following 

conditions is true. 

(i) At least 75 percent of the days in each quarter of each of three consecutive years have at least 

one reported hourly value, and the design value calculated according to the procedures specified 

in section 5.2 is above the level of the primary or secondary 1-hour standard.

(ii) (A) A 1-hour primary or secondary standard design value that is below the level of the 

NAAQS can be validated if the substitution test in section 3.2(c)(ii)(B) results in a “test design 

value” that is below the level of the NAAQS. The test substitutes actual ‘‘high’’ reported daily 

maximum 1-hour values from the same site at about the same time of the year (specifically, in 

the same calendar quarter) for unknown values that were not successfully measured. Note that 

the test is merely diagnostic in nature, intended to confirm that there is a very high likelihood 

that the original design value (the one with less than 75 percent data capture of hours by day and 

of days by quarter) reflects the true under-NAAQS-level status for that 3-year period; the result 

of this data substitution test (the ‘‘test design value”, as defined in section 3.2(c)(ii)(B)) is not 

considered the actual design value. For this test, substitution is permitted only if there are at least 

200 days across the three matching quarters of the three years under consideration (which is 

about 75 percent of all possible daily values in those three quarters) for which 75 percent of the 

hours in the day, including state-flagged data affected by exceptional events which have been 
Page 228 of 238

approved for exclusion by the Administrator, have reported concentrations. However, maximum 

1-hour values from days with less than 75 percent of the hours reported shall also be considered 

in identifying the high value to be used for substitution.

(B) The substitution test is as follows: Data substitution will be performed in all quarter periods 

that have less than 75 percent data capture but at least 50 percent data capture, including stateflagged data affected by exceptional events which have been approved for exclusion by the 

Administrator; if any quarter has less than 50 percent data capture then this substitution test 

cannot be used. Identify for each quarter (e.g., January-March) the highest reported daily

maximum 1-hour value for that quarter, excluding state-flagged data affected by exceptional 

events which have been approved for exclusion by the Administrator, looking across those three 

months of all three years under consideration. All daily maximum 1-hour values from all days in 

the quarter period shall be considered when identifying this highest value, including days with 

less than 75 percent data capture. If after substituting the highest non-excluded reported daily 

maximum 1-hour value for a quarter for as much of the missing daily data in the matching 

deficient quarter(s) as is needed to make them 100 percent complete, the procedure in section 5.2 

yields a recalculated 3-year 1-hour standard “test design value” below the level of the standard, 

then the 1-hour primary or secondary standard design value is deemed to have passed the 

diagnostic test and is valid, and the level of the standard is deemed to have been met in that 3-

year period. As noted in section 3.2(c)(i), in such a case, the 3-year design value based on the 

data actually reported, not the ‘‘test design value’’, shall be used as the valid design value. 

(iii) (A) A 1-hour primary or secondary standard design value that is above the level of the 

NAAQS can be validated if the substitution test in section 3.2(c)(iii)(B) results in a “test design 

value” that is above the level of the NAAQS. The test substitutes actual ‘‘low” reported daily 
Page 229 of 238

maximum 1-hour values from the same site at about the same time of the year (specifically, in 

the same three months of the calendar) for unknown values that were not successfully measured. 

Note that the test is merely diagnostic in nature, intended to confirm that there is a very high 

likelihood that the original design value (the one with less than 75 percent data capture of hours 

by day and of days by quarter) reflects the true above-NAAQS-level status for that 3-year period; 

the result of this data substitution test (the ‘‘test design value”, as defined in section 

3.2(c)(iii)(B)) is not considered the actual design value. For this test, substitution is permitted 

only if there are a minimum number of available daily data points from which to identify the low 

quarter-specific daily maximum 1-hour values, specifically if there are at least 200 days across 

the three matching quarters of the three years under consideration (which is about 75 percent of 

all possible daily values in those three quarters) for which 75 percent of the hours in the day have 

reported concentrations. Only days with at least 75 percent of the hours reported shall be 

considered in identifying the low value to be used for substitution.

(B) The substitution test is as follows: Data substitution will be performed in all quarter periods 

that have less than 75 percent data capture. Identify for each quarter (e.g., January-March) the 

lowest reported daily maximum 1-hour value for that quarter, looking across those three months 

of all three years under consideration. All daily maximum 1-hour values from all days with at 

least 75 percent capture in the quarter period shall be considered when identifying this lowest 

value. If after substituting the lowest reported daily maximum 1-hour value for a quarter for as 

much of the missing daily data in the matching deficient quarter(s) as is needed to make them 75 

percent complete, the procedure in section 5.2 yields a recalculated 3-year 1-hour standard “test 

design value” above the level of the standard, then the 1-hour primary or secondary standard 

design value is deemed to have passed the diagnostic test and is valid, and the level of the 
Page 230 of 238

standard is deemed to have been exceeded in that 3-year period. As noted in section 3.2(c)(i), in 

such a case, the 3-year design value based on the data actually reported, not the ‘‘test design 

value’’, shall be used as the valid design value.

(d) A 1-hour primary or secondary standard design value based on data that do not meet the 

completeness criteria stated in 3.2(b) and also do not satisfy section 3.2(c), may also be 

considered valid with the approval of, or at the initiative of, the Administrator, who may consider 

factors such as monitoring site closures/moves, monitoring diligence, the consistency and levels 

of the valid concentration measurements that are available, and nearby concentrations in 

determining whether to use such data.

(e) The procedures for calculating the 1-hour primary and secondary standard design values are 

given in section 5.2 of this appendix.

4. Rounding Conventions. 

4.1 Rounding Conventions for the Annual Primary and Secondary NO2 NAAQS.

* * * * *

(b) The annual primary or secondary standard design value is calculated pursuant to section 5.1 

and then rounded to the nearest whole number or 1 ppb (decimals 0.5 and greater are rounded up 

to the nearest whole number, and any decimal lower than 0.5 is rounded down to the nearest 

whole number).

4.2 Rounding Conventions for the 1-hour Primary and Secondary NO2 NAAQS.

* * * * *

(c) The 1-hour primary or secondary standard design value is calculated pursuant to section 5.2 

and then rounded to the nearest whole number or 1 ppb (decimals 0.5 and greater are rounded up 
Page 231 of 238

to the nearest whole number, and any decimal lower than 0.5 is rounded down to the nearest 

whole number).

5. Calculation Procedures for the Primary and Secondary NO2 NAAQS.

5.1 Procedures for the Annual Primary and Secondary NO2 NAAQS

* * * * *

(b) The annual primary or secondary standard design value for a site is the valid annual mean 

rounded according to the conventions in section 4.1.

5.2 Calculation Procedures for the 1-hour Primary and Secondary NO2 NAAQS.

* * * * *

(b) The 1-hour primary or secondary standard design value for a site is the mean of the three 

annual 98th percentile values, rounded according to the conventions in section 4.

* * * * *

5. Appendix T is amended as follows:

a. by revising paragraph 1.(a),

b. by revising the definition of “Design values” under paragraph 1.(c),

c. by revising paragraph 2.(b),

d. by revising paragraphs 3.(a) through (e),

e. by revising paragraph 4.(c), and

f. by revising paragraph 5.(b) to read as follows:
Page 232 of 238

Appendix T to Part 50—Interpretation of the Primary and Secondary National Ambient 

Air Quality Standards for Oxides of Sulfur (Sulfur Dioxide) 

1. General.

(a) This appendix explains the data handling conventions and computations necessary for 

determining when the primary and secondary national ambient air quality standards for Oxides 

of Sulfur as measured by Sulfur Dioxide (“SO2 NAAQS”) specified in § 50.17 and §50.5 (b), 

respectively, are met at an ambient air quality monitoring site. Sulfur Dioxide (SO2) is measured 

in the ambient air by a Federal reference method (FRM) based on appendix A-1 or A-2 to this 

part or by a Federal equivalent method (FEM) designated in accordance with part 53 of this 

chapter. Data handling and computation procedures to be used in making comparisons between 

reported SO2 concentrations and the levels of the SO2 NAAQS are specified in the following 

sections.

* * * * *

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

* * * * *

Design values are the metrics (i.e., statistics) that are compared to the NAAQS levels to 

determine compliance, calculated as specified in section 5 of this appendix. The design value for 

the primary and secondary 1-hour NAAQS is the 3-year average of annual 99th percentile daily 

maximum 1-hour values for a monitoring site (referred to as the “1-hour primary standard design 

value”).
Page 233 of 238

* * * * *

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

Reporting Considerations.

* * * * *

(b) Data from two or more monitors from the same year at the same site reported to EPA under 

distinct Pollutant Occurrence Codes shall not be combined in an attempt to meet data 

completeness requirements. The Administrator will combine annual 99th percentile daily 

maximum concentration values from different monitors in different years, selected as described 

here, for the purpose of developing a valid 1-hour primary or secondary standard design value. 

If more than one of the monitors meets the completeness requirement for all four quarters of a 

year, the steps specified in section 5(a) of this appendix shall be applied to the data from the 

monitor with the highest average of the four quarterly completeness values to derive a valid 

annual 99th percentile daily maximum concentration. If no monitor is complete for all four 

quarters in a year, the steps specified in section 3(c) and 5(a) of this appendix shall be applied to 

the data from the monitor with the highest average of the four quarterly completeness values in 

an attempt to derive a valid annual 99th percentile daily maximum concentration. This paragraph 

does not prohibit a monitoring agency from making a local designation of one physical monitor 

as the primary monitor for a Pollutant Occurrence Code and substituting the 1-hour data from a 

second physical monitor whenever a valid concentration value is not obtained from the primary

monitor; if a monitoring agency substitutes data in this manner, each substituted value must be 

accompanied by an AQS qualifier code indicating that substitution with a value from a second 

physical monitor has taken place.
Page 234 of 238

* * * * *

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

(a) The 1-hour primary or secondary SO2 NAAQS is met at an ambient air quality monitoring 

site when the valid 1-hour primary or secondary standard design value is less than or equal to 75 

parts per billion (ppb).

(b) An SO2 1-hour primary or secondary standard design value is valid if it encompasses three 

consecutive calendar years of complete data. A year meets data completeness requirements 

when all 4 quarters are complete. A quarter is complete when at least 75 percent of the sampling 

days for each quarter have complete data. A sampling day has complete data if 75 percent of the 

hourly concentration values, including State-flagged data affected by exceptional events which 

have been approved for exclusion by the Administrator, are reported.

(c) In the case of one, two, or three years that do not meet the completeness requirements of 

section 3(b) of this appendix and thus would normally not be useable for the calculation of a 

valid 3-year 1-hour primary or secondary standard design value, the 3-year 1-hour primary or 

secondary standard design value shall nevertheless be considered valid if one of the following 

conditions is true. 

(i) At least 75 percent of the days in each quarter of each of three consecutive years have at least 

one reported hourly value, and the design value calculated according to the procedures specified 

in section 5 is above the level of the primary or secondary 1-hour standard.

(ii) (A) A 1-hour primary or secondary standard design value that is equal to or below the level 

of the NAAQS can be validated if the substitution test in section 3(c)(ii)(B) results in a “test 

design value” that is below the level of the NAAQS. The test substitutes actual ‘‘high’’ reported 

daily maximum 1-hour values from the same site at about the same time of the year (specifically, 
Page 235 of 238

in the same calendar quarter) for unknown values that were not successfully measured. Note that 

the test is merely diagnostic in nature, intended to confirm that there is a very high likelihood 

that the original design value (the one with less than 75 percent data capture of hours by day and 

of days by quarter) reflects the true under-NAAQS-level status for that 3-year period; the result 

of this data substitution test (the ‘‘test design value”, as defined in section 3(c)(ii)(B)) is not 

considered the actual design value. For this test, substitution is permitted only if there are at least 

200 days across the three matching quarters of the three years under consideration (which is 

about 75 percent of all possible daily values in those three quarters) for which 75 percent of the 

hours in the day, including State-flagged data affected by exceptional events which have been 

approved for exclusion by the Administrator, have reported concentrations. However, maximum 

1-hour values from days with less than 75 percent of the hours reported shall also be considered 

in identifying the high value to be used for substitution.

(B) The substitution test is as follows: Data substitution will be performed in all quarter periods 

that have less than 75 percent data capture but at least 50 percent data capture, including Stateflagged data affected by exceptional events which have been approved for exclusion by the 

Administrator; if any quarter has less than 50 percent data capture then this substitution test 

cannot be used. Identify for each quarter (e.g., January-March) the highest reported daily

maximum 1-hour value for that quarter, excluding State-flagged data affected by exceptional 

events which have been approved for exclusion by the Administrator, looking across those three 

months of all three years under consideration. All daily maximum 1-hour values from all days in 

the quarter period shall be considered when identifying this highest value, including days with 

less than 75 percent data capture. If after substituting the highest reported daily maximum 1-hour 

value for a quarter for as much of the missing daily data in the matching deficient quarter(s) as is 
Page 236 of 238

needed to make them 100 percent complete, the procedure in section 5 yields a recalculated 3-

year 1-hour standard “test design value” less than or equal to the level of the standard, then the 1-

hour primary or secondary standard design value is deemed to have passed the diagnostic test 

and is valid, and the level of the standard is deemed to have been met in that 3-year period. As 

noted in section 3(c)(i), in such a case, the 3-year design value based on the data actually

reported, not the ‘‘test design value’’, shall be used as the valid design value.

(iii) (A) A 1-hour primary or secondary standard design value that is above the level of the 

NAAQS can be validated if the substitution test in section 3(c)(iii)(B) results in a “test design 

value” that is above the level of the NAAQS. The test substitutes actual ‘‘low” reported daily 

maximum 1-hour values from the same site at about the same time of the year (specifically, in 

the same three months of the calendar) for unknown hourly values that were not successfully

measured. Note that the test is merely diagnostic in nature, intended to confirm that there is a 

very high likelihood that the original design value (the one with less than 75 percent data capture 

of hours by day and of days by quarter) reflects the true above-NAAQS-level status for that 3-

year period; the result of this data substitution test (the ‘‘test design value”, as defined in section 

3(c)(iii)(B)) is not considered the actual design value. For this test, substitution is permitted only

if there are a minimum number of available daily data points from which to identify the low 

quarter-specific daily maximum 1-hour values, specifically if there are at least 200 days across 

the three matching quarters of the three years under consideration (which is about 75 percent of 

all possible daily values in those three quarters) for which 75 percent of the hours in the day have 

reported concentrations. Only days with at least 75 percent of the hours reported shall be 

considered in identifying the low value to be used for substitution.
Page 237 of 238

(B) The substitution test is as follows: Data substitution will be performed in all quarter periods 

that have less than 75 percent data capture. Identify for each quarter (e.g., January-March) the 

lowest reported daily maximum 1-hour value for that quarter, looking across those three months 

of all three years under consideration. All daily maximum 1-hour values from all days with at 

least 75 percent capture in the quarter period shall be considered when identifying this lowest 

value. If after substituting the lowest reported daily maximum 1-hour value for a quarter for as 

much of the missing daily data in the matching deficient quarter(s) as is needed to make them 75 

percent complete, the procedure in section 5 yields a recalculated 3-year 1-hour standard “test 

design value” above the level of the standard, then the 1-hour primary or secondary standard 

design value is deemed to have passed the diagnostic test and is valid, and the level of the 

standard is deemed to have been exceeded in that 3-year period. As noted in section 3(c)(i), in 

such a case, the 3-year design value based on the data actually reported, not the ‘‘test design 

value’’, shall be used as the valid design value.

(d) A 1-hour primary or secondary standard design value based on data that do not meet the 

completeness criteria stated in 3(b) and also do not satisfy section 3(c), may also be considered 

valid with the approval of, or at the initiative of, the Administrator, who may consider factors 

such as monitoring site closures/moves, monitoring diligence, the consistency and levels of the 

valid concentration measurements that are available, and nearby concentrations in determining 

whether to use such data.

(e) The procedures for calculating the 1-hour primary or secondary standard design values are 

given in section 5 of this appendix.

4. Rounding Conventions for the 1-hour Primary and Secondary SO2 NAAQS. 
Page 238 of 238

* * * * * 

(c) The 1-hour primary or secondary standard design value is calculated pursuant to section 5 

and then rounded to the nearest whole number or 1 ppb (decimals 0.5 and greater are rounded up 

to the nearest whole number, and any decimal lower than 0.5 is rounded down to the nearest 

whole number).

5. Calculation Procedures for the 1-hour Primary and Secondary SO2 NAAQS.

* * * * *

(b) The 1-hour primary or secondary standard design value for an ambient air quality monitoring 

site is the mean of the three annual 99th percentile values, rounded according to the conventions 

in section 4.