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1 ENVIRONMENTAL PROTECTION AGENCY

2 40 CFR Part 50

3 [EPA-HQ-OAR-2007-1145]

4 RIN: 2060-AO72

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

6

7 AGENCY: Environmental Protection Agency (EPA).

8 ACTION: Proposed rule.

9 SUMMARY: Based on its review of the air quality criteria and the secondary national ambient 

10 air quality standards (NAAQS) for oxides of nitrogen and oxides of sulfur, EPA proposes to 

11 retain the current nitrogen dioxide (NO2) and sulfur dioxide (SO2) secondary standards to 

12 provide requisite protection for the direct effects on vegetation resulting from exposure to 

13 gaseous oxides of nitrogen and sulfur in the ambient air. With regard to protection from the 

14 deposition of oxides of nitrogen and sulfur to sensitive aquatic and terrestrial ecosystems, 

15 including acidification and nutrient enrichment effects, EPA is proposing to add secondary

16 standards identical to the NO2 and SO2 primary 1-hour standards and not to set a new multi17 pollutant secondary standard in this review. In addition, EPA has decided to undertake a field 

18 pilot program to gather and analyze additional relevant data so as to enhance the Agency’s 

19 understanding of the degree of protectiveness that a new multi-pollutant approach, defined in 

20 terms of an aquatic acidification index (AAI), would afford and to support development of an 

21 appropriate monitoring network for such a standard. The EPA solicits comment on the 

22 framework of such a standard and on the design of the field pilot program.
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1 DATES: Written comments on this proposed rule must be received by [insert date 60 days 

2 after date of publication in the Federal Register].

3 Public Hearings: The EPA intends to hold a public hearing around the end of August to 

4 early September and will announce in a separate Federal Register notice the date, time, and 

5 address of the public hearing on this proposed rule.

6 ADDRESSES: Submit your comments, identified by Docket ID No. EPA-HQ-OAR-2007-

7 1145, by one of the following methods:

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

9  Email: a-and-r-Docket@epa.gov.

10  Fax: 202-566-1741.

11  Mail: Docket No. EPA-HQ-OAR-2007-1145, Environmental Protection Agency, Mail 

12 code 6102T, 1200 Pennsylvania Ave., NW., Washington, DC 20460. Please include a 

13 total of two copies.

14  Hand Delivery: Docket No. EPA-HQ-OAR-2007-1145, Environmental Protection 

15 Agency, EPA West, Room 3334, 1301 Constitution Ave., NW, Washington, DC. Such 

16 deliveries are only accepted during the Docket’s normal hours of operation, and special 

17 arrangements should be made for deliveries of boxed information.

18 Instructions: Direct your comments to Docket ID No. EPA-HQ-OAR-2007-1145. The 

19 EPA’s policy is that all comments received will be included in the public docket without change 

20 and may be made available online at www.regulations.gov, including any personal information 

21 provided, unless the comment includes information claimed to be Confidential Business 

22 Information (CBI) or other information whose disclosure is restricted by statute. Do not submit 

23 information that you consider to be CBI or otherwise protected through www.regulations.gov or 
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1 email. The www.regulations.gov website is an “anonymous access” system, which means EPA 

2 will not know your identity or contact information unless you provide it in the body of your 

3 comment. If you send an email comment directly to EPA without going through 

4 www.regulations.gov, your email address will be automatically captured and included as part of 

5 the comment that is placed in the public docket and made available on the Internet. If you 

6 submit an electronic comment, EPA recommends that you include your name and other contact 

7 information in the body of your comment and with any disk or CD-ROM you submit. If EPA 

8 cannot read your comment due to technical difficulties and cannot contact you for clarification, 

9 EPA may not be able to consider your comment. Electronic files should avoid the use of special 

10 characters, any form of encryption, and be free of any defects or viruses. For additional 

11 information about EPA’s public docket, visit the EPA Docket Center homepage at 

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

13 Docket: All documents in the docket are listed in the www.regulations.gov index. 

14 Although listed in the index, some information is not publicly available, e.g., CBI or other 

15 information whose disclosure is restricted by statute. Certain other material, such as copyrighted 

16 material, will be publicly available only in hard copy. Publicly available docket materials are 

17 available either electronically in www.regulations.gov or in hard copy at the Air and Radiation 

18 Docket and Information Center, EPA/DC, EPA West, Room 3334, 1301 Constitution Ave., NW, 

19 Washington, DC. The Public Reading Room is open from 8:30 a.m. to 4:30 p.m., Monday 

20 through Friday, excluding legal holidays. The telephone number for the Public Reading Room is 

21 (202) 566-1744 and the telephone number for the Air and Radiation Docket and Information 

22 Center is (202) 566-1742. 
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1 FOR FURTHER INFORMATION CONTACT: Dr. Richard Scheffe, Office of Air Quality

2 Planning and Standards, U.S. Environmental Protection Agency, Mail code C304-02, Research 

3 Triangle Park, NC 27711; telephone: 919-541-4650; fax: 919-541-2357; email: 

4 scheffe.rich@epa.gov. 

5 SUPPLEMENTARY INFORMATION:

6 General Information

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

8 1. Submitting CBI. Do not submit this information to EPA through www.regulations.gov or 

9 email. Clearly mark the part or all of the information that you claim to be CBI. For CBI 

10 information in a disk or CD ROM that you mail to EPA, mark the outside of the disk or 

11 CD ROM as CBI and then identify electronically within the disk or CD ROM the specific 

12 information that is claimed as CBI. In addition to one complete version of the comment 

13 that includes information claimed as CBI, a copy of the comment that does not contain 

14 the information claimed as CBI must be submitted for inclusion in the public docket. 

15 Information so marked will not be disclosed except in accordance with procedures set 

16 forth in 40 CFR part 2.

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

18  Identify the rulemaking by docket number and other identifying information (subject 

19 heading, Federal Register date and page number).

20  Follow directions – The Agency may ask you to respond to specific questions or 

21 organize comments by referencing a Code of Federal Regulations (CFR) part or section 

22 number.
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1  Explain why you agree or disagree, suggest alternatives, and substitute language for your 

2 requested changes.

3  Describe any assumptions and provide any technical information and/or data that you 

4 used.

5  If you estimate potential costs or burdens, explain how you arrived at your estimate in 

6 sufficient detail to allow for it to be reproduced.

7  Provide specific examples to illustrate your concerns, and suggest alternatives.

8  Explain your views as clearly as possible.Make sure to submit your comments by the 

9 comment period deadline identified.

10 Availability of Related Information

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

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

13 REPORT (ISA) is available on EPAs National Center for Environmental Assessment web site. 

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

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

16 and other related technical documents are available on EPA’s Office of Air Quality Planning and 

17 Standards (OAQPS) Technology Transfer Network (TTN) web site. The Policy Assessment is 

18 available at http://www.epa.gov/ttn/naaqs/standards/no2so2sec/cr_pa.html, and the exposure and 

19 risk assessments and other related technical documents are available at 

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

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

22 Table of Contents

23 The following topics are discussed in this preamble:
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1 I. Background

2 A. Legislative Requirements

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

4 1. NAAQS for Oxides of Nitrogen

5 2. NAAQS for Oxides of Sulfur

6 C. History of Related Assessments and Agency Actions

7 D. History of the Current Review

8 E. Scope of the Current Review

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

10 A. Ecological Effects

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

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

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

14 2. Acidification Effects Associated with Deposition of Oxides of Nitrogen 

15 and Sulfur

16 a. Nature of Acidification –related Ecosystem Responses

17 i. Aquatic ecosystems

18 ii. Terrestrial ecosystems

19 iii. Ecosystem sensitivity

20 b. Magnitude of Acidification-related Ecosystem Responses

21 i. Aquatic acidification

22 ii. Terrestrial acidification

23 c. Key Uncertainties Associated with Acidification

24 i. Aquatic acidification

25 ii. Terrestrial acidification

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

27 Nitrogen 

28 a. Nature of Nutrient Enrichment-related Ecosystem Responses

29 i. Aquatic ecosystems

30 ii. Terrestrial ecosystems

31 iii. Ecosystem sensitivity to nutrient enrichment

32 b. Magnitude of Nutrient Enrichment-related Ecosystem Responses

33 i. Aquatic ecosystems

34 ii. Terrestrial ecosystems

35 c. Key Uncertainties Associated with Nutrient Enrichment

36 i. Aquatic ecosystems

37 ii. Terrestrial ecosystems

38 4. Other Ecological Effects

39 B. Risk and Exposure Assessment

40 1. Overview of Risk and Exposure Assessment

41 2. Key Findings

42 a. Air quality analyses

43 b. Deposition-related aquatic acidification

44 c. Deposition-related terrestrial acidification

45 d. Deposition-related aquatic nutrient enrichment

46 e. Deposition-related terrestrial nutrient enrichment 
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1 f. Additional effects

2 3. Conclusions on Effects

3 C. Adversity of Effects to Public Welfare

4 1. Ecosystem Services

5 2. Effects on Ecosystem Services

6 a. Aquatic Acidification

7 b. Terrestrial Acidification

8 c. Nutrient Enrichment

9 3. Summary

10 D. Adequacy of the Current Standards

11 1. Adequacy of the Current Standards for Direct Effects

12 2. Appropriateness and Adequacy of the Current Standards for Deposition13 related Effects

14 a. Appropriateness

15 b. Adequacy of Protection

16 i. Aquatic Acidification

17 ii. Terrestrial acidification

18 iii. Terrestrial nutrient enrichment

19 iv. Aquatic nutrient enrichment

20 v. Other effects

21 3. CASAC Views

22 4. Administrator’s Proposed Conclusions Concerning Adequacy of Current 

23 Standard

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

25 Standards for Aquatic Acidification

26 A. Ambient Air Indicators

27 1. Oxides of Sulfur

28 2. Oxides of Nitrogen

29 B. Form

30 1. Ecological Indicator

31 2. Linking ANC to Deposition

32 3. Linking Deposition to Ambient Air Indicators

33 4. Aquatic Acidification Index

34 5. Spatial Aggregation

35 a. Ecoregion Sensitivity

36 b. Representative Ecoregion-specific Factors

37 i. Factor F1

38 a) Acid-sensitive Ecoregions

39 b) Non-acid sensitive Ecoregions

40 ii. Factor F2

41 iii. Factors F3 and F4

42 c. Factors in Data-limited Ecoregions

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

44 6. Summary of the AAI Form

45 C. Averaging Time

46 D. Level
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1 1. Association Between pH Levels and Target ANC Levels 

2 2. ANC Levels Related to Effects on Aquatic Ecosystems 

3 3. Consideration of Episodic Acidity

4 4. Consideration of Ecosystem Response Time 

5 5. Prior Examples of Target ANC Levels 

6 6. Consideration of Public Welfare Benefits

7 7. Summary of Alternative Levels

8 E. Combined Alternative Levels and Forms

9 F. Characterization of Uncertainties

10 1. Overview of Uncertainty

11 2. Uncertainties Associated with Data Gaps

12 3. Uncertainties in Modeled Processes

13 G. CASAC Advice

14 H. Administrator’s Proposed Conclusions

15 IV. Field Pilot Program and Ambient Monitoring 

16 A. Field Pilot Program

17 a. Objectives

18 b. Overview of Field Pilot Program

19 c. Complementary Measurements

20 d. Complementary Research Efforts

21 e. Implementation Challenges

22 f. Final Monitoring Plan Development and Stakeholder Participation

23 B. Evaluation of Monitoring Methods

24 1. Potential FRMs for SO2 and p-SO4

25 2. Potential FRM for NOy

26 V. Statutory and Executive Order Reviews

27 A. Executive Order 12866: Regulatory Planning and Review

28 B. Paperwork Reduction Act

29 C. Regulatory Flexibility Act

30 D. Unfunded Mandates Reform Act

31 E. Executive Order 13132: Federalism

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

33 Governments

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

35 Safety Risks

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

37 Distribution, or Use

38 I. National Technology Transfer and Advancement Act

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

40 Minority Populations and Low-Income Populations

41 References

42
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1 I. Background

2 A. Legislative Requirements

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

19 criteria, is requisite to protect the public welfare from any known or anticipated adverse effects 

20 associated with the presence of [the] pollutant in the ambient air.” Welfare effects as defined in 

21 section 302(h) (42 U.S.C. § 7602(h)) include, but are not limited to, “effects on soils, water, 

 

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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1 crops, vegetation, man-made materials, animals, wildlife, weather, visibility and climate, damage 

2 to and deterioration of property, and hazards to transportation, as well as effects on economic 

3 values and on personal comfort and well-being.”

4 In setting standards that are “requisite” to protect public health and welfare, as provided 

5 in section 109(b), EPA’s task is to establish standards that are neither more nor less stringent 

6 than necessary for these purposes. In so doing, EPA may not consider the costs of implementing 

7 the standards. See generally, Whitman v. American Trucking Associations, 531 U.S. 457, 465-

8 472, 475-76 (2001). Likewise, “[a]ttainability and technological feasibility are not relevant 

9 considerations in the promulgation of national ambient air quality standards.” American 

10 Petroleum Institute v. Costle, 665 F. 2d at 1185. Section 109(d)(1) requires that “not later than 

11 December 31, 1980, and at 5-year intervals thereafter, the Administrator shall complete a 

12 thorough review of the criteria published under section 108 and the national ambient air quality

13 standards . . . and shall make such revisions in such criteria and standards and promulgate such 

14 new standards as may be appropriate . . . .” Section 109(d)(2) requires that an independent 

15 scientific review committee “shall complete a review of the criteria . . . and the national primary

16 and secondary ambient air quality standards . . . and shall recommend to the Administrator any

17 new . . . standards and revisions of existing criteria and standards as may be appropriate . . . .” 

18 Since the early 1980's, this independent review function has been performed by the Clean Air 

19 Scientific Advisory Committee (CASAC).

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

21 1. NAAQS for Oxides of Nitrogen

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

23 with oxides of nitrogen, EPA promulgated identical primary and secondary NAAQS for NO2 in 
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1 April 1971. These standards were set at a level of 0.053 parts per million (ppm) as an annual 

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

3 EPA, 1982), which updated the scientific criteria upon which the initial standards were based. In 

4 February 1984 EPA proposed to retain these standards (49 FR 6866). After taking into account 

5 public comments, EPA published the final decision to retain these standards in June 1985 (50 FR 

6 25532).

7 The EPA reviewed the oxides of nitrogen secondary standards in 1987. In November 

8 1991 EPA released an updated draft air quality criteria document (AQCD) for CASAC and 

9 public review and comment (56 FR 59285), which provided a comprehensive assessment of the 

10 available scientific and technical information on health and welfare effects associated with NO2 

11 and other oxides of nitrogen. The CASAC reviewed the draft document at a meeting held on 

12 July 1, 1993 and concluded in a closure letter to the Administrator that the document “provides a 

13 scientifically balanced and defensible summary of current knowledge of the effects of this 

14 pollutant and provides an adequate basis for EPA to make a decision as to the appropriate 

15 NAAQS for NO2” (Wolff, 1993). The Air Quality Criteria for Oxides of Nitrogen was then 

16 finalized (US EPA, 1995a). EPA’s OAQPS also prepared a Staff Paper that summarized and 

17 integrated the key studies and scientific evidence contained in the revised AQCD for oxides of 

18 nitrogen and identified the critical elements to be considered in the review of the NO2 NAAQS. 

19 CASAC reviewed two drafts of the Staff Paper and concluded in a closure letter to the 

20 Administrator that the document provided a “scientifically adequate basis for regulatory

21 decisions on nitrogen dioxide” (Wolff, 1995).

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

23 the primary or secondary NAAQS for NO2 (60 FR 52874; October 11, 1995). A year later, the 
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1 Administrator made a final determination not to revise the NAAQS for NO2 after careful 

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

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

4 standard with the establishment of a new 1-hour standard (75 FR 6474), the secondary NAAQS 

5 for NO2 remains 0.053 ppm (100 micrograms per cubic meter [μg/m3] of air), annual arithmetic 

6 average, calculated as the arithmetic mean of the 1-hour NO2 concentrations.

7 2. NAAQS for Oxides of Sulfur

8 EPA promulgated primary and secondary NAAQS for SO2 in April 1971 (36 FR 8186). 

9 The secondary standards included a standard set at 0.02 ppm, annual arithmetic mean, and a 3-

10 hour average standard set at 0.5 ppm, not to be exceeded more than once per year. These 

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

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

13 Vegetation”) of Air Quality Criteria for Sulfur Oxides (US EPA, 1973) indicated that it could not 

14 properly be concluded that the vegetation injury reported resulted from the average SO2 

15 exposure over the growing season, rather than from short-term peak concentrations. Therefore, 

16 EPA proposed (38 FR 11355) and then finalized (38 FR 25678) a revocation of the annual mean 

17 secondary standard. At that time, EPA was aware that then-current concentrations of oxides of 

18 sulfur in the ambient air had other public welfare effects, including effects on materials, 

19 visibility, soils, and water. However, the available data were considered insufficient to establish 

20 a quantitative relationship between specific ambient concentrations of oxides of sulfur and such 

21 public welfare effects (38 FR 25679).

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

23 with that for particulate matter (PM) and would produce a combined PM and oxides of sulfur 
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1 criteria document. Following its review of a draft revised criteria document in August 1980, 

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

18 Papers, Volumes I and II (US EPA, 1984a, b) and The Acidic Deposition Phenomenon and Its 

19 Effects: Critical Assessment Document (US EPA, 1985) (53 FR 14935 -14936). These 

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

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

22 deposition completed by EPA at that point.
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1 In April 1988 (53 FR 14926), EPA proposed not to revise the existing primary and 

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

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

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

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

6 been decreased through ongoing research efforts, EPA would draft and support an appropriate set 

7 of control measures. Although EPA revised the primary SO2 standard in June 2010 by

8 establishing a new 1-hour standard and revoking the existing 24-hour and annual standards (75 

9 FR 35520), no further decisions on the secondary SO2 standard have been published. 

10 C. History of Related Assessments and Agency Actions

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

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

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

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

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

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

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

18 acid deposition problem, it soon became clear that a program to address acid deposition was 

19 needed. The Clean Air Act Amendments of 1990 included numerous separate provisions related 

20 to the acid deposition problem. The primary and most important of the provisions, the 

21 amendments to Title IV of the Act, established the Acid Rain Program to reduce emissions of 

22 SO2 by 10 million tons and emissions of nitrogen oxides by 2 million tons from 1980 emission 

23 levels in order to achieve reductions over broad geographic regions. In this provision, Congress 
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1 included a statement of findings that led them to take action, concluding that (1) the presence of 

2 acid compounds and their precursors in the atmosphere and in deposition from the atmosphere 

3 represents a threat to natural resources, ecosystems, materials, visibility, and public health; (2) 

4 the problem of acid deposition is of national and international significance; and (3) current and 

5 future generations of Americans will be adversely affected by delaying measures to remedy the 

6 problem. 

7 Second, Congress authorized the continuation of the NAPAP in order to assure that the 

8 research and monitoring efforts already undertaken would continue to be coordinated and would 

9 provide the basis for an impartial assessment of the effectiveness of the Title IV program.

10 Third, Congress considered that further action might be necessary in the long term to 

11 address any problems remaining after implementation of the Title IV program and, reserving 

12 judgment on the form that action could take, included Section 404 of the 1990 Amendments 

13 (Clean Air Act Amendments of 1990, Pub. L. 101-549, § 404) requiring EPA to conduct a study

14 on the feasibility and effectiveness of an acid deposition standard or standards to protect 

15 “sensitive and critically sensitive aquatic and terrestrial resources.” At the conclusion of the 

16 study, EPA was to submit a report to Congress. Five years later, EPA submitted its report, 

17 entitled Acid Deposition Standard Feasibility Study: Report to Congress (US EPA, 1995b) in 

18 fulfillment of this requirement. That Report concluded that establishing acid deposition 

19 standards for sulfur and nitrogen deposition may at some point in the future be technically

20 feasible, although appropriate deposition loads for these acidifying chemicals could not be 

21 defined with reasonable certainty at that time. 

22 Fourth, the 1990 Amendments also added new language to sections of the CAA 

23 pertaining to the scope and application of the secondary NAAQS designed to protect the public 
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1 welfare. Specifically, the definition of “effects on welfare” in Section 302(h) was expanded to 

2 state that the welfare effects include effects “...whether caused by transformation, conversion, or 

3 combination with other air pollutants.” 

4 In 1999, seven Northeastern states cited this amended language in Section 302(h) in a 

5 petition asking EPA to use its authority under the NAAQS program to promulgate secondary

6 NAAQS for the criteria pollutants associated with the formation of acid rain. The petition stated 

7 that this language “clearly references the transformation of pollutants resulting in the inevitable 

8 formation of sulfate and nitrate aerosols and/or their ultimate environmental impacts as wet and 

9 dry deposition, clearly signaling Congressional intent that the welfare damage occasioned by 

10 sulfur and nitrogen oxides be addressed through the secondary standard provisions of Section 

11 109 of the Act.” The petition further stated that “recent federal studies, including the NAPAP 

12 Biennial Report to Congress: An Integrated Assessment, document the continued and increasing 

13 damage being inflicted by acid deposition to the lakes and forests of New York, New England 

14 and other parts of our nation, demonstrating that the Title IV program had proven insufficient.” 

15 The petition also listed other adverse welfare effects associated with the transformation of these 

16 criteria pollutants, including impaired visibility, eutrophication of coastal estuaries, global 

17 warming, and tropospheric ozone and stratospheric ozone depletion.

18 In a related matter, the Office of the Secretary of the U.S. Department of Interior (DOI) 

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

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

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

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

23 waters; visibility impairment; and foliar injury from ozone.
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1 In a Federal Register notice in 2001, EPA announced receipt of these requests and asked 

2 for comment on the issues raised in them. EPA stated that it would consider any relevant 

3 comments and information submitted, along with the information provided by the petitioners and 

4 DOI, before making any decision concerning a response to these requests for rulemaking (65 FR 

5 48699).

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

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

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

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

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

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

12 2005).

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

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

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

16 of the appropriateness of setting secondary NAAQS to address remaining known or anticipated 

17 adverse public welfare effects resulting from the acidic and nutrient deposition of these criteria 

18 pollutants.

19 D. History of the Current Review

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

21 FR 73236) for the development of a revised Integrated Science Assessment for Oxides of 

22 Nitrogen and Oxides of Sulfur- Ecological Criteria (henceforth the "ISA"). An Integrated 

23 Review Plan (IRP) was developed to provide the framework and schedule as well as the scope of
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1 the review and to identify policy-relevant questions to be addressed in the components of the 

2 review. The IRP was released in 2007 (US EPA, 2007) for CASAC and public review. EPA 

3 held a workshop in July 2007 on the ISA to obtain broad input from the relevant scientific 

4 communities. This workshop helped to inform the preparation of the first draft ISA, which was 

5 released for CASAC and public review in December 2007; a CASAC meeting was held on April 

6 2-3, 2008 to review the first draft ISA. A second draft ISA was released for CASAC and public 

7 review in August 2008, and was discussed at a CASAC meeting held on October 1-2, 2008. The

8 final ISA (US EPA, 2008) was released in December 2008. 

9 Based on the science presented in the ISA, EPA developed a Risk and Exposure 

10 Assessment for Review of the Secondary National Ambient Air Quality Standards for Oxides of 

11 Nitrogen and Oxides of Sulfur (henceforth the “REA”) to further assess the national impact of 

12 the effects documented in the ISA. The Draft Scope and Methods Plan for Risk/Exposure 

13 Assessment: Secondary NAAQS Review for Oxides of Nitrogen and Oxides of Sulfur outlining 

14 the scope and design of the future REA was prepared for CASAC and public in March 2008. A 

15 first draft REA was presented to CASAC and the public for review in August 2008 and a second 

16 draft was presented for review in June 2009. The final REA (US EPA 2009) was released in 

17 September 2009. A first draft Policy Assessment was released in March 2010 and reviewed by

18 CASAC on April 1-2, 2010. In a June 22, 2010 letter to the Administrator, CASAC provided 

19 advice and recommendations to the Agency concerning the first draft Policy Assessment (Russell 

20 and Samet, 2010a). A second draft Policy Assessment was released to CASAC and the public 

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

22 and recommendations to the Agency regarding the second draft Policy Assessment in a 

23 December 9, 2010 letter (Russell and Samet 2010b). CASAC and public comments on the 
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1 second draft Policy Assessment were considered by EPA staff in developing a final Policy

2 Assessment (US EPA, 2011). CASAC requested an additional meeting to provide additional 

3 advice to the Administrator based on the final Policy Assessment on February 15–16, 2011. 

4 EPA released on January 14, 2011, the final Policy Assessment prior to final document 

5 production, to provide sufficient time for CASAC review of the document in advance of this 

6 meeting. A final Policy Assessment, incorporating final reference checks and document 

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

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

9 revisions to the secondary NAAQS for oxides of nitrogen and oxides of sulfur. 

10 On February 3, 2006, the Center for Biological Diversity and four other plaintiffs filed 

11 an amended complaint alleging that EPA had failed to complete the current review within the 

period provided by statute.212 The schedule for completion of this review is governed by a consent 

13 decree resolving that lawsuit and the subsequent extension agreed to by the parties. The schedule 

14 presented in the original consent decree that governs this review, entered by the court on 

15 November 19, 2007, was revised on October 22, 2009 to allow for a 17 month extension of the 

16 schedule. The current decree provides that EPA sign for publication notices of proposed and 

17 final rulemaking concerning its review of the oxides of nitrogen and oxides of sulfur NAAQS no 

18 later than July 12, 2011 and March 20, 2012, respectively. 

19 This action presents the Administrator’s proposed decisions on the review of the current 

20 secondary oxides of nitrogen and oxides of sulfur standards. Throughout this preamble a number 

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

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

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

 

2 Center for Biological Diversity, et al. v. Johnson, No. 05–1814 (D.D.C.)
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1 comments on all issues involved with this proposal, including all such proposed judgments, 

2 conclusions, findings, and determinations.

3 E. Scope of the Current Review

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

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

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

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

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

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

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

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

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

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

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

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

16 these interactions and associated ecological effects.

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

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

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

20 atmosphere, and under certain conditions, can combine with atmospheric ammonia to form 

21 ammonium nitrate, a component of fine PM. Oxides of sulfur are precursors to the formation of 

22 particulate sulfate, which is a significant component of fine PM in many parts of the U.S. There 

23 are a number of welfare effects directly associated with ozone and fine PM, including ozone-
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1 related damage to vegetation and PM-related visibility impairment. Protection against those 

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

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

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

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

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

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

8 deposition in the form of particulate nitrate and particulate sulfate.

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

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

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

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

13 ecosystems. Non-managed terrestrial ecosystems tend to have a higher fraction of N deposition 

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

15 that agricultural and commercial forest lands are routinely fertilized with amounts of nitrogen 

16 that exceed air pollutant inputs even in the most polluted areas (US EPA, 2008, section 3.3.9). 

17 This review recognizes that the effects of nitrogen deposition in managed areas are viewed 

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

19 natural, unmanaged ecosystems, largely due to the more homogeneous, controlled nature of 

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

21 increased productivity in those ecosystems.

22 In focusing on natural sensitive ecosystems, the Policy Assessment primarily considers 

23 the effects of ambient oxides of nitrogen and sulfur via deposition on multiple ecological 
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1 receptors. The ISA highlights effects including those associated with acidification and nitrogen 

2 nutrient enrichment. With a focus on these deposition-related effects, EPA’s objective is to 

3 develop a framework for oxides of nitrogen and sulfur standards that incorporates ecologically 

4 relevant factors and that recognizes the interactions between the two pollutants as they deposit to 

5 sensitive ecosystems. The overarching policy objective is to develop a secondary standard(s)

6 that is based on the ecological criteria described in the ISA and the results of the assessments in 

7 the REA, and is consistent with the requirement of the CAA to set secondary standards that are 

8 requisite to protect the public welfare from any known or anticipated adverse effects associated 

9 with the presence of these air pollutants in the ambient air. Also consistent with the CAA, this 

10 policy objective necessarily includes consideration of “variable factors . . . which of themselves 

11 or in combination with other factors may alter the effects on public welfare” of the criteria air 

12 pollutants included in this review.

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

14 sulfur on ecological impacts on sensitive aquatic ecosystems associated with acidifying 

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

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

17 and advice from CASAC on earlier drafts of this Policy Assessment (Russell and Samet, 2010a, 

18 2010b), and as discussed in detail in the Policy Assessment, we have the greatest confidence in 

19 the causal linkages between oxides of nitrogen and sulfur and aquatic acidification effects 

20 relative to other deposition-related effects, including terrestrial acidification and aquatic and 

21 terrestrial nutrient enrichment.

22
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1 II. Rationale for Proposed Decision on the Adequacy of the Current Secondary 

2 Standards

3 Decisions on retaining or revising the current secondary standards for oxides of nitrogen 

4 and sulfur are largely public welfare policy judgments based on the Administrator’s informed 

5 assessment of what constitutes requisite protection against adverse effects to public welfare. A 

6 public welfare policy decision should draw upon scientific information and analyses about 

7 welfare effects, exposure and risks, as well as judgments about the appropriate response to the 

8 range of uncertainties that are inherent in the scientific evidence and analyses. The ultimate 

9 determination as to what level of damage to ecosystems and the services provided by those 

10 ecosystems is adverse to public welfare is not wholly a scientific question, although it is 

11 informed by scientific studies linking ecosystem damage to losses in ecosystem services, and 

12 information on the value of those losses of ecosystem services. In reaching such decisions, the 

13 Administrator seeks to establish standards that are neither more nor less stringent than necessary

14 for this purpose.

15 This section presents the rationale for the Administrator’s proposed conclusions with 

16 regard to the adequacy of protection and ecological relevance of the current secondary standards 

17 for oxides of nitrogen and sulfur. As discussed more fully below, this rationale considered the 

18 latest scientific information on ecological effects associated with the presence of oxides of 

19 nitrogen and oxides of sulfur in the ambient air. This rationale also takes into account: (1) staff 

20 assessments of the most policy-relevant information in the ISA and staff analyses of air quality, 

21 exposure, and ecological risks, presented more fully in the REA and in the Policy Assessment,

22 upon which staff conclusions on revisions to the secondary oxides of nitrogen and oxides of 

23 sulfur standards are based; (2) CASAC advice and recommendations, as reflected in discussions 
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1 of drafts of the ISA, REA, and Policy Assessment at public meetings, in separate written 

2 comments, and in CASAC’s letters to the Administrator; and (3) public comments received 

3 during the development of these documents, either in connection with CASAC meetings or 

4 separately. 

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

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

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

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

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

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

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

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

13 and/or sulfur contribute to the reported effects.

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

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

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

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

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

19 the ecological indicator for a targeted effect.

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

21 an ecologically meaningful secondary standard in the Policy Assessment and to further develop 

22 the indicator, form, level, and averaging time of such as standard in section III of this proposal is 
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1 depicted below and highlights the three key linkages that need to be considered in developing an 

2 ecologically relevant standard.

3

4

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

6 for oxides of nitrogen and sulfur.

7

8 The following discussion relies heavily on Chapters 2 and 3 of the Policy Assessment. 

9 The Policy Assessment includes staff’s evaluation of the policy implications of the scientific 

10 assessment of the evidence presented and assessed in the ISA and the results of quantitative 

11 assessments based on that information presented and assessed in the REA. Taken together, this 

12 information informs staff conclusions and the development of policy options in the Policy

13 Assessment for consideration in addressing public and welfare effects associated with the 

14 presence of oxides of nitrogen and oxides of sulfur in the ambient air. Of particular note, chapter 

15 2 of the Policy Assessment presents information not repeated here that characterizes emissions, 

16 air quality, deposition and water quality. It includes discussions of the sources of nitrogen and 

17 sulfur in the atmosphere as well as current ambient air quality monitoring networks and models. 

18 Additional information in this section includes ecological modeling and water quality data 

19 sources. 

20 Section II.A presents a discussion of the effects associated with oxides of nitrogen and 

21 sulfur in the ambient air. The discussion is organized around the types of effects being 
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1 considered, including direct effects of gaseous oxides of nitrogen and sulfur, deposition-related 

2 effects related to acidification and nutrient enrichment, and other effects such as materials 

3 damage, climate-related effects and mercury methylation.

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

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

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

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

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

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

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

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

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

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

14 linked to ecological indicators.

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

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

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

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

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

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

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

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

23 and whether they are adequate to protect against such effects is presented.
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1 A. Ecological Effects

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

3 of nitrogen and sulfur, including the direct effects of gas-phase exposure to oxides of nitrogen 

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

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

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

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

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

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

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

11 evidence of such effects are also discussed throughout this section. 

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

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

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

15 pollutants such as sulfur dioxide (SO2), nitrogen dioxide (NO2), nitric oxide (NO), nitric acid 

16 (HNO3) and peroxyacetyl nitrite (PAN) are associated with negative impacts to vegetation. The 

17 current secondary NAAQS were set to protect against direct damage to vegetation by exposure to 

18 gas-phase oxides of nitrogen and sulfur, such as foliar injury, decreased photosynthesis, and 

19 decreased growth. The following summary is a concise overview of the known or anticipated 

20 effects to vegetation caused by gas phase nitrogen and sulfur. Most phototoxic effects associated 

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

22 observed in the U.S. (US EPA, 2008, section 3.4.2.4).
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1 a. Nature of ecosystem responses to gas-phase nitrogen and sulfur

2 The 2008 ISA found that gas phase N and S are associated with direct phytotoxic effects 

3 (US EPA, 2008, section 4.4). The evidence is sufficient to infer a causal relationship between 

4 exposure to SO2 and injury to vegetation (US EPA, 2008, section 4.4.1 and 3.4.2.1). Acute foliar 

5 injury to vegetation from SO2 may occur at levels above the current secondary standard (3-h 

6 average of 0.50 ppm). Effects on growth, reduced photosynthesis and decreased yield of 

7 vegetation are also associated with increased SO2 exposure concentration and time of exposure.

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

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

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

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

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

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

14 sensitive lichen species. 

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

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

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

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

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

20 3.4.2). The functional relationship between ambient concentrations of gas phase oxides of 

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

22 stomatal conductance and plant detoxification mechanisms, and external factors including plant 
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1 water status, light, temperature, humidity, and pollutant exposure regime (US EPA, 2008, section

2 3.4.2).

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

4 as well as to stomatal aperture. The aperture of the stomata is controlled largely by the 

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

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

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

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

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

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

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

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

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

14 nitrogen fixation, decreased K+ absorption, and structural changes.

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

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

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

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

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

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

21 and HNO3 at concentrations currently observed in the United States with the exception of some 

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

23 contributing to the decline in lichen species in the Los Angeles basin. Most phototoxic effects 
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1 associated with gas phase oxides of nitrogen and sulfur occur at levels well above ambient 

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

3 2. Acidification Effects Associated with Deposition of Oxides of Nitrogen and Sulfur

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

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

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

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

8 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-

9

) and nitrate (NO3

-

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

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

12 occurring in the soils and waterbodies (US EPA, 2009, section 2.1). The chemical forms of 

13 nitrogen that may contribute to acidifying deposition include both oxidized and reduced chemical 

14 species, including NHx.

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

2-

or NO3

-

15 , 

16 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+ 17 ), 

magnesium (Mg2+), sodium (Na+), or potassium (K+), rather than hydrogen (H+18 ) and dissolved 

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

decreases. Continued SO4

2-

or NO3

-

20 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

-

21 (with 

H

+

and Al3+ 22 ) leads to acidification of soil water, and by connection, surface water. Introduction 

23 of strong acid anions such as sulfate and nitrate to an already acidic soil, whether naturally or due 
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1 to anthropogenic activities, can lead to instantaneous acidification of waterbodies through direct 

2 runoff without any significant change in base cation saturation. The ability of a watershed to 

3 neutralize acidic deposition is determined by a variety of biogeophysical factors including 

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

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

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

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

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

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

10 years to decades result in increases in inputs of N and S to ecosystems and the associated 

11 ecological effects. Episodic or short term (i.e., hours or days) deposition refers to events in which 

12 the level of the acid neutralizing capacity (ANC) of a lake or stream is temporarily lowered. In 

13 aquatic ecosystems, short-term (i.e., hours or days) episodic changes in water chemistry can have 

14 significant biological effects. Episodic acidification refers to conditions during precipitation or 

15 snowmelt events when proportionately more drainage water is routed through upper soil horizons 

16 that tend to provide less acid neutralizing than is passing through deeper soil horizons (US EPA, 

17 2009, section 4.2). In addition, the accumulated sulfate and nitrate in snow packs can provide a 

18 surge of acidic inputs. Some streams and lakes may have chronic or base flow chemistry that is 

19 suitable for aquatic biota, but may be subject to occasional acidic episodes with deleterious 

20 consequences to sensitive biota.

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

22 by acidification to ecosystems within the United States. Acidification affects both terrestrial and 

23 freshwater aquatic ecosystems. 
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1 a. Nature of acidification-related ecosystem responses

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

3 varying degrees of acidification of ecosystems (US EPA, 2008). In the process of acidification, 

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

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

6 through the soil and eventually leaches into adjacent water bodies.

7 i. Aquatic ecosystems

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

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

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

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

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

13 ANC is a key indicator of acidification with relevance to both terrestrial and aquatic ecosystems. 

14 ANC is useful because it integrates the overall acid-base status of a lake or stream and reflects 

15 how aquatic ecosystems respond to acidic deposition over time. There is also a relationship 

16 between ANC and the surface water constituents that directly contribute to or ameliorate acidityrelated stress, in particular, concentrations of hydrogen ion (as pH), Ca2+ 17 , and aluminum. 

18 Moreover, low pH surface waters leach aluminum from soils, which is quite lethal to fish and 

19 other aquatic organisms. In aquatic systems, there is a direct relationship between ANC and fish 

20 and phyto-zooplankton diversity and abundance. 

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

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

23 species richness is positively correlated with pH and ANC because energy cost in maintaining 
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1 physiological homeostasis, growth, and reproduction is high at low ANC levels. For example, 

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

3 Study Area lakes that indicates the probability of occurrence of an organism for a given value of 

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

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

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

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

8 also begins to decline. When ANC concentrations are <50 μeq/L, they are generally associated 

9 with death or loss of fitness of biota that are sensitive to acidification.

10 Consistent and coherent documentation from multiple studies on various species from all 

11 major trophic levels of aquatic systems shows that geochemical alteration caused by acidification 

12 can result in the loss of acid-sensitive biological species (US EPA, 2008, section 3.2.3.3). This is 

13 most often discussed with relation to pH. For example, in the Adirondacks, of the 53 fish species 

14 recorded in Adirondack lakes about half (26 species) were absent from lakes with pH below 6.0. 

15 Biological effects are linked to changes in water chemistry including decreases in ANC and pH 

16 and increases in inorganic Al concentration. The direct biological effects are caused by lowered 

17 pH which leads to increased inorganic Al concentrations (US EPA, 2011, Figures 3-1 and 3-2). 

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

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

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

21 which are summarized in section 3.1.1 of the Policy Assessment. Significant harm to sensitive 

22 aquatic species has been observed at pH levels below 6. Normal stream pH levels with little to 

23 no toxicity range from 6 to 7 (MacAvoy et al, 1995). Baker et al (1990) observed that “lakes 
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1 with pH less than approximately 6.0 contain significantly fewer species than lakes with pH levels 

2 above 6.0”. As noted in Chapter 3, typically at pH <4.5 and an ANC <0 μeq/L, complete to 

3 near-complete loss of many taxa of organisms occur, including fish and aquatic insect 

4 populations, whereas other taxa are reduced to only acidophilic species. 

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

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

7 observed in the Shenandoah National Park. In field observations, when pH levels dropped to 5, 

8 mortality rates went to 100 percent. (Bulger et al, 2000). At pH levels ranging from 5.4 to 5.8, 

9 cumulative mortality continues to increase. Several studies have shown that trout exposed to 

10 water with varying pH levels and fish larvae showed increasing mortality as pH levels decrease. 

11 In one study almost 100 percent mortality was observed at a pH of 4.5 compared to almost 100 

12 percent survival at a pH of 6.5. Intermediate pH values (6.0, 5.5) in all cases showed reduced 

13 survival compared with the control (6.5), but not by statistically significant amounts (US EPA, 

14 2008, section 3.2.3.3). 

15 One important indicator of acid stress is increased fish mortality. The response of fish to 

16 pH is not uniform across species. A number of synoptic surveys indicated loss of species 

17 diversity and absence of several fish species in the pH range of 5.0 to 5.5. If pH is lower, there is 

18 a greater likelihood that more fish species could be lost without replacement, resulting in 

19 decreased richness and diversity. In general, populations of salmonids are not found at pH levels 

20 less than 5.0, and smallmouth bass (Micropterus dolomieu) populations are usually not found at 

21 pH values less than about 5.2 to 5.5. From Table 3-1, only one study showed significant 

22 mortality effects above a pH of 6, while a number of studies showed significant mortality when 

23 pH levels are at or below 5.5. 
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1 The highest pH level for any of the studies reported in the ISA, is 6.0, suggesting that pH 

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

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

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

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

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

7 protection against mortality effects throughout the life stages of fish.

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

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

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

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

12 section 3.2.3.3).

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

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

15 53 potential species). There is often a positive relationship between pH and number of fish 

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

17 100 μeq/L. Such observed relationships are complicated, however, by the tendency for smaller 

18 lakes and streams, having smaller watersheds, to also support fewer fish species, irrespective of 

19 acid-base chemistry. This pattern may be due to a decrease in the number of available niches as 

20 stream or lake size decreases. Nevertheless, fish species richness is relatively easily determined 

21 and is one of the most useful indicators of biological effects of surface water acidification. 

22 Changes in stream water pH and ANC also contribute to declines in taxonomic richness

23 of zooplankton, and macroinvertebrates which are often sources of food for fish, birds and other 
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1 animal species in various ecosystems. These fish may also serve as a source of food and 

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

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

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

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

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

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

8 Adirondacks. 

9 ii. Terrestrial ecosystems

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

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

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

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

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

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

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

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

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

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

20 effect can be neutralized by weathering from geologic parent material or base cation exchange. 

The Ca2+ 21 and Al concentrations in soil water are strongly influenced by soil acidification and 

22 both have been shown to have quantitative links to tree health, including Al interference with 

Ca2+ 23 uptake and Al toxicity to roots. Effects of nitrification and associated acidification and 
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1 cation leaching have been consistently shown to occur only in soils with a C:N ratio below about 

2 20 to 25.

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

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

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

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

7 increased over the past 40 years, a period that coincided with increased emissions of S and N 

8 oxides and increased acidifying deposition. Acidifying deposition can contribute to dieback in 

9 sugar maple (Acer saccharum) through depletion of cations from soil with low levels of available 

10 Ca. Grasslands are likely less sensitive to acidification than forests due to grassland soils being 

11 generally rich in base cations.

12 iii. Ecosystem sensitivity 

13 The intersection between current deposition loading, historic loading, and sensitivity

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

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

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

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

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

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

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

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

22 base-poor bedrock, high relief, and shallow soils (US EPA, 2008, section 3.2.4.1).
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1 b. Magnitude of acidification-related ecosystem responses

2 Terrestrial and aquatic ecosystems differ in their response to acidifying deposition. 

3 Therefore the magnitude of ecosystem response is described separately for aquatic and terrestrial 

4 ecosystems in the following sections. The magnitude of response refers to both the severity of 

5 effects and the spatial extent of the U.S. which is affected.

6 i. Aquatic acidification

7 Freshwater ecosystem surveys and monitoring in the eastern United States have been 

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

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

10 Integrated Monitoring of Ecosystems (TIME), and Long-term Monitoring (LTM) programs. 

11 Based on analyses of surface water data from these programs, New England, the Adirondack 

12 Mountains, the Appalachian Mountains (northern Appalachian Plateau and Ridge/Blue Ridge 

13 region), and the Upper Midwest contain the most sensitive lakes and streams (i.e., ANC less than 

14 about 50 μeq/L). Portions of northern Florida also contain many acidic and low-ANC lakes and 

15 streams, although the role of acidifying deposition in this region is less clear. The western U.S. 

16 contains many of the surface waters most sensitive to potential acidification effects, but with the 

17 exception of the Los Angeles Basin and surrounding areas, the levels of acidifying deposition are 

18 low in most areas. Therefore, acidification of surface waters by acidic deposition is not as 

19 prevalent in the western U.S., and the extent of chronic surface water acidification that has 

20 occurred in that region to date has likely been very limited relative to the Eastern U.S. (US EPA, 

21 2008, section 3.2.4.2 and US EPA, 2009, section 4.2.2).

22 There are a number of species including fish, aquatic insects, other invertebrates and 

23 algae that are sensitive to acidification and cannot survive, compete, or reproduce in acidic 
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1 waters (US EPA, 2008, section 3.2.3.3). Decreases in ANC and pH have been shown to 

2 contribute to declines in species richness and declines in abundance of zooplankton, 

3 macroinvertebrates, and fish. Reduced growth rates have been attributed to acid stress in a 

4 number of fish species including Atlantic salmon (Salmo salar), Chinook salmon (Oncorhynchus 

5 tshawytscha), lake trout (Salvelinus namaycush), rainbow trout (Oncorhynchis mykiss), brook 

6 trout (Salvelinus Fontinalis), and brown trout (Salmo trutta). In response to small to moderate 

7 changes in acidity, acid-sensitive species are often replaced by other more acid-tolerant species, 

8 resulting in changes in community composition and richness. The effects of acidification are 

9 continuous, with more species being affected at higher degrees of acidification. At a point, 

10 typically a pH <4.5 and an ANC <0 μeq/L, complete to near-complete loss of many taxa of 

11 organisms occur, including fish and aquatic insect populations, whereas other taxa are reduced to 

12 only acidophilic species. These changes in taxa composition are associated with the high energy

13 cost in maintaining physiological homeostasis, growth, and reproduction at low ANC levels (US 

14 EPA, 2008, section 3.2.3.3). Decreases in species richness related to acidification have been 

15 observed in the Adirondack Mountains and Catskill Mountains of New York, New England and 

16 Pennsylvania, and Virginia. From the sensitive areas identified by the ISA, further “case study” 

17 analyses on aquatic ecosystems in the Adirondack Mountains and Shenandoah National Park 

18 were conducted to better characterize ecological risk associated with acidification (US EPA, 

19 2009, section 4).

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

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

22 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

-

23 concentrations and ANC levels were 
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1 analyzed to affirm the understanding that reductions in deposition could influence the risk of 

2 acidification. ANC values have been categorized according to their effects on biota, as shown in 

3 the table below. Monitoring data from TIME/LTM and EMAP programs were assessed for the 

4 years 1990 to 2006, and past, present, and future water quality levels were estimated by both 

5 steady-state and dynamic biogeochemical models.
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1

2 Table II-1. Ecological effects associated with alternative levels of acid neutralizing capacity 

3 (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.

4

5 Studies on fish species richness in the Adirondacks Case Study Area demonstrated the 

6 effect of acidification. Of the 53 fish species recorded in Adirondack Case Study Area lakes, 

7 only 27 species were found in lakes with a pH <6.0. The 26 species missing from lakes with a 

8 pH <6.0 include important recreational species, such as Atlantic salmon, tiger trout (Salmo trutta
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1 X Salvelinus fontinalis), redbreast sunfish (Lepomis auritus), bluegill (Lepomis macrochirus), 

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

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

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

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

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

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

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

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

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

declines in NO3

-

and SO4

2-

11 concentrations in surface waters. The 2006 concentrations are still 

12 above pre-acidification (1860) conditions. MAGIC modeling predicts surface water 

concentrations of NO3

-

and SO4

2-

13 are10- and 32-fold higher, respectively, in 2006 than in 1860. 

14 The estimated average ANC across 60 streams in the Shenandoah Case Study Area is 57.9 μeq/L 

15 (± 4.5 μeq/L). 55% of all monitored streams in the Shenandoah Case Study Area have a current 

16 risk of Elevated, Severe, or Acute. Of the 55%, 18% are chronically acidic today (US EPA, 

17 2009, section 4.2.4.3).

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

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

20 Elevated to Acute problems with acidity in 2050. In fact, from 2006 to 2050, the percentage of 

21 streams with Acute Concern is predicted to increase by 5%, while the percentage of streams in 

22 Moderate Concern decreases by 5%.
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1 Biological effects of increased acidification documented in the Shenandoah Case Study

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

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

4 lower by one fish species for every 21 μeq/L decrease in ANC in Shenandoah National Park 

5 streams (US EPA, 2008, section 3.2.3.4).

6 ii. Terrestrial acidification

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

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

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

(such as Ca2+, Mg2+ 10 and potassium), as well as soil Al content. Tree species show a range of 

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

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

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

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

15 0.2 to 2.5 (US EPA, 2009).

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

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

18 However, one preliminary national evaluation estimated that ~15% of forest ecosystems in the 

19 U.S. exceed the estimated critical load based on soil ANC leaching for S and N deposition by

20 >250 eq ha-1 yr-1 (McNulty et al., 2007). Forests of the Adirondack Mountains of New York, 

21 Green Mountains of Vermont, White Mountains of New Hampshire, the Allegheny Plateau of 

22 Pennsylvania, and high-elevation forest ecosystems in the southern Appalachians are the regions 

23 most sensitive to terrestrial acidification effects from acidifying deposition (US EPA, 2008, 
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1 section 3.2.4.2). While studies show some recovery of surface waters, there are widespread 

2 measurements of ongoing depletion of exchangeable base cations in forest soils in the 

3 northeastern U.S. despite recent decreases in acidifying deposition, indicating a slow recovery

4 time.

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

6 in the eastern United States by using Bc/Al ratio in acidified forest soils as an indicator to assess 

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

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

9 spruce growth can be decreased by 20%. Sugar maple growth can be decreased by 20% at a 

10 Bc/Al ratio of 0.6 (US EPA, 2009, section 4.4). The REA analysis determined the health of at 

11 least a portion of the sugar maple and red spruce growing in the United States may have been 

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

13 sulfur deposition levels exceeded three selected critical loads for tree growth in 3% to 75% of all 

14 sugar maple plots across 24 states--that is, it exceeded the highest (least stringent) of the three 

15 critical loads in 3% of plots, and the lowest (most stringent) in 75% of plots. For red spruce, total 

16 nitrogen and sulfur deposition levels exceeded three selected critical loads in 3% to 36% of all 

17 red spruce plots across eight states (US EPA, 2009, section 4.4). 

18 c. Key uncertainties associated with acidification

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

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

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

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

23 relationship between the selected ecological indicator and atmospheric deposition as described 
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1 by the ecological effect function (US EPA, 2009, Figure 1-1). A further discussion of 

2 uncertainty in aquatic and terrestrial ecosystems is presented below. The key uncertainties were 

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

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

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

6 robustness of data sets, monitoring networks, availability of data that allows for extrapolation 

7 to larger assessment areas, and input parameters for modeling and developing the ecological 

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

9 into this criterion.

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

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

12 by the scientific community for their application in this analysis.

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

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

15 atmospheric deposition and the ecological indicator of an effect.

16 i. Aquatic acidification

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

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

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

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

21 1990 to 2006 used for this analysis as well as EPA EMAP and REMAP surveys, provide 

22 considerable data on surface water trends. 
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1 There is fairly high confidence associated with modeling and input parameters. 

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

3 site calibrations. The 95% confidence interval for pre-acidification of lakes was an average of 15 

4 μeq/L difference in ANC concentrations or 10% and 8 μeq/L or 5% for streams (US EPA, 2009, 

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

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

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

8 2007). To quantify uncertainty in the model parameters, Monte Carlo methods were employed 

9 to develop an inverse function of exceedances. There is high confidence associated with the 

10 ecological effect function developed for aquatic acidification. In calculating the ANC function, 

11 the depositional load for N or S is fixed by the deposition of the other, so deposition for either 

12 will never be zero (US EPA, 2009, Figure 7.1-6).

13 ii. Terrestrial acidification 

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

15 robust and considered high quality. The USFS-Kane Experimental Forest and significant 

16 amounts of research work in the Allegheny Plateau have produced extensive, peer-reviewed data 

17 sets. A meta-analysis of laboratory studies showed that tree growth was decreased by 20% 

18 relative to controls for BC/Al ratios (US EPA, 2008, section 7.2.1 and Figure 7.2-1). Sugar 

19 maple and red spruce were the focus of the REA since they are demonstrated to be negatively 

affected by soil available Ca2+ 20 depletion and high concentrations of available Al, and occur in 

21 areas that receive high acidifying deposition, There is high confidence about the use of the REA 

22 terrestrial acidification data and their value for extrapolating to a larger regional population of 

23 forests. 
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1 There is high confidence associated with the models, input parameters, and assessment of 

2 uncertainty used in the case study for terrestrial acidification. The Simple Mass Balance (SMB) 

3 model, a commonly used and widely applied approach for estimating critical loads, was used in 

4 the REA analysis (US EPA, 2008, section 7.2.2). There is fairly high confidence associated with 

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

6 7.2.3).

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

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

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

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

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

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

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

14 other sources of N greatly exceed N deposition. Both oxides of nitrogen and reduced forms of 

15 nitrogen (NHx) contribute to nitrogen deposition. For the most part, nitrogen effects on 

16 ecosystems do not depend on whether the nitrogen is in oxidized or reduced form. Thus, this 

17 summary focuses on the effects of nitrogen deposition in total. 

18 a. Nature of nutrient enrichment-related ecosystem responses

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

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

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

22 that leads to effects on biological organisms. 

23 i. Aquatic ecosystems
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1 In freshwater ecosystems, the evidence is sufficient to infer a causal relationship between 

2 N deposition and the alteration of biogeochemical cycling in freshwater aquatic ecosystems (US 

3 EPA, 2008, section 3.3.2.3). N deposition is the main source of N enrichment to headwater 

4 streams, lower order streams and high elevation lakes. The most common chemical indicators 

that were studied included NO3

−

5 and dissolved inorganic nitrogen (DIN) concentration in surface 

waters as well as Chl a:total P ratio. Elevated surface water NO3

−

6 concentrations occur in both 

7 the eastern and western U.S. Studies report a significant correlation between N deposition and

8 lake biogeochemistry by identifying a correlation between wet deposition and [DIN] and Chl a: 

9 Total P. Recent evidence provides examples of lakes and streams that are limited by N and show 

10 signs of eutrophication in response to N addition.

11 The evidence is sufficient to infer a causal relationship between N deposition and the 

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

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

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

15 In the ISA, the evidence is sufficient to infer a causal relationship between N deposition 

16 and the biogeochemical cycling of N and carbon (C) in estuaries (US EPA, 2008, section 4.3.4.1 

17 and 3.3.2.3). In general, estuaries tend to be nitrogen-limited, and many currently receive high 

18 levels of nitrogen input from human activities (US EPA, 2009, section 5.1.1). It is unknown if 

19 atmospheric deposition alone is sufficient to cause eutrophication; however, the contribution of 

20 atmospheric nitrogen deposition to total nitrogen load is calculated for some estuaries and can be 

21 >40% (US EPA, 2009, section 5.1.1).

22 The evidence is sufficient to infer a causal relationship between N deposition and the 

23 alteration of species richness, species composition and biodiversity in estuarine ecosystems (US 
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1 EPA, 2008, section 4.3.4.2 and 3.3.5.4). Atmospheric and non-atmospheric sources of N 

2 contribute to increased phytoplankton and algal productivity, leading to eutrophication. Shifts in 

3 community composition, reduced hypolimnetic DO, decreases in biodiversity, and mortality of 

4 submerged aquatic vegetation are associated with increased N deposition in estuarine systems. 

5 ii. Terrestrial Ecosystems

6 The evidence is sufficient to infer a causal relationship between N deposition and the 

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

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

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

10 of NO3- in stream water were identified as two of the primary indictors of N enrichment. Several 

11 N-addition studies indicate that NO3- leaching is induced by chronic additions of N. Studies 

12 identified in the ISA found that surface water NO3- concentrations exceeded 1 μeq/L in 

13 watersheds receiving about 9 to 13 kg N/ha/yr of atmospheric N deposition. N deposition 

14 disrupts the nutrient balance of ecosystems with numerous biogeochemical effects. The chemical 

indicators that are typically measured include NO3

−

15 leaching, soil C:N ratio, rates of N 

16 mineralization, nitrification, denitrification, foliar N concentration, and soil water NO3 − and 

NH4

+

17 concentrations. Note that N saturation (N leaching from ecosystems) does not need to 

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

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

20 Due to the complexity of interactions between the N and C cycling, the effects of N on C budgets 

21 (quantified input and output of C to the ecosystem) are variable. Regional trends in net 

22 ecosystem productivity (NEP) of forests (not managed for silviculture) have been estimated 

23 through models based on gradient studies and meta-analysis. Atmospheric N deposition has been 
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1 shown to cause increased litter accumulation and carbon storage in above-ground woody 

2 biomass. In the West, this has lead to increased susceptibility to more severe fires. Less is 

3 known regarding the effects of N deposition on C budgets of non-forest ecosystems.

4 The evidence is sufficient to infer a causal relationship between N deposition on the 

5 alteration of species richness, species composition and biodiversity in terrestrial ecosystems (US 

6 EPA, 2008, section 4.3.1.2). Some organisms and ecosystems are more sensitive to N deposition 

7 and effects of N deposition are not observed in all habitats. The most sensitive terrestrial taxa to 

8 N deposition are lichens. Empirical evidence indicates that lichens in the U.S. are affected by

9 deposition levels as low as 3 kg N/ha/yr. Alpine ecosystems are also sensitive to N deposition, 

10 changes in an individual species (Carex rupestris) were estimated to occur at deposition levels 

11 near 4 kg N /ha/yr and modeling indicates that deposition levels near 10 kg N/ha/yr alter plant 

12 community assemblages. In several grassland ecosystems, reduced species diversity and an 

13 increase in non-native, invasive species are associated with N deposition. 

14 iii. Ecosystem sensitivity to nutrient enrichment

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

16 sensitivity to N deposition (US EPA, 2008, Table 4-4). Increased deposition to N-limited 

17 ecosystems can lead to production increases that may be either beneficial or adverse depending 

18 on the system and management goals. 

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

20 nutrient availability. Change in the availability of one important nutrient, such as N, may result 

21 in an imbalance in ecological stoichiometry, with effects on ecosystem processes, structure and 

22 function. In general, N deposition to terrestrial ecosystems causes accelerated growth rates in 

23 some species deemed desirable in commercial forests but may lead to altered competitive 
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1 interactions among species and nutrient imbalances, ultimately affecting biodiversity. The onset 

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

3 ecosystems to N deposition. In aquatic ecosystems, N that is both leached from the soil and 

4 directly deposited to the water surface can pollute the surface water. This causes alteration of the 

5 diatom community at levels as low as 1.5 kg N/ha/yr in sensitive freshwater ecosystems. 

6 The degree of ecosystem effects lies at the intersection of N loading and N-sensitivity. 

7 N-sensitivity is predominately driven by the degree to which growth is limited by nitrogen 

8 availability. Grasslands in the western United States are typically N-limited ecosystems 

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

10 the ISA (US EPA, 2008, section 3.3.3), indicated that N fertilization increased aboveground 

11 growth in all non-forest ecosystems except for deserts. In other words, almost all terrestrial 

12 ecosystems are N-limited and will be altered by the addition of anthropogenic nitrogen. 

13 Likewise, a freshwater lake or stream must be N-limited to be sensitive to N-mediated 

14 eutrophication. There are many examples of fresh waters that are N-limited or N and 

15 phosphorous (P) co-limited (US EPA, 2008, section 3.3.3.2). A large dataset meta-analysis 

16 discussed in the ISA (US EPA, 2008, section 3.3.3.2), found that N-limitation occurred as 

17 frequently as P-limitation in freshwater ecosystems. Additional factors that govern the 

18 sensitivity of ecosystems to nutrient enrichment from N deposition include rates and form of N 

19 deposition, elevation, climate, species composition, plant growth rate, length of growing season, 

20 and soil N retention capacity (US EPA, 2008, section 4.3). Less is known about the extent and 

21 distribution of the terrestrial ecosystems in the U.S. that are most sensitive to the effects of 

22 nutrient enrichment from atmospheric N deposition compared to acidification.
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1 Because the productivity of estuarine and near shore marine ecosystems is generally 

2 limited by the availability of N, they are susceptible to the eutrophication effect of N deposition 

3 (US EPA, 2008, section 4.3.4.1). A recent national assessment of eutrophic conditions in 

4 estuaries found the most eutrophic estuaries were generally those that had large watershed-to5 estuarine surface area, high human population density, high rainfall and runoff, low dilution, and 

6 low flushing rates. In the REA, the National Oceanic and Atmospheric Administration’s 

7 (NOAA) National Estuarine Eutrophication Assessment (NEEA) assessment tool, Assessment of 

8 Estuarine Tropic Status (ASSETS) categorical Eutrophication Index (EI) was used to evaluate 

9 eutrophication due to atmospheric loading of nitrogen. ASSETS EI is an estimation of the 

10 likelihood that an estuary is experiencing eutrophication or will experience eutrophication based 

11 on five ecological indicators: chlorophyll a, macroalgae, dissolved oxygen, nuisance/toxic algal 

12 blooms and submerged aquatic vegetation (SAV). 

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

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

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

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

17 results were then considered with SPAtially Referenced Regression (SPARROW) modeling to 

18 develop a response curve to examine the role of atmospheric nitrogen deposition in achieving a 

19 desired decrease in load. To change the Neuse River Estuary’s EI score from Bad to Poor not 

20 only must 100% of the total atmospheric nitrogen deposition be eliminated, but considerably 

21 more nitrogen from other sources as well must be controlled (US EPA, 2009, section 5.2.7.2). In 

22 the Potomac River estuary, a 78% decrease of total nitrogen could move the EI score from Bad

23 to Poor (US EPA, 2009, section 5.2.7.1). The results of this analysis indicated decreases in 
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1 atmospheric deposition alone could not eliminate coastal eutrophication problems due to 

2 multiple non-atmospheric nitrogen inputs (US EPA, 2009, section 7.3.3). However, the 

3 somewhat arbitrary discreteness of the EI scale can mask the benefits of decreases in nitrogen 

4 between categories.

5 In general, estuaries tend to be N-limited, and many currently receive high levels of N 

6 input from human activities to cause eutrophication. As reported in the ISA (US EPA, 2008, 

7 section 3.2.2.2), atmospheric N loads to estuaries in the U.S. are estimated to range from 2-8% 

8 for Guadalupe Bay, TX on the lowest end to as high as 72% for St Catherines-Sapelo estuary, 

9 GA. The Chesapeake Bay is an example of a large, well-studied and severely eutrophic estuary

10 that is calculated to receive as much as 30% of its total N load from the atmosphere.

11 b. Magnitude of ecosystem responses

12 i. Aquatic ecosystems

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

14 conditions and how ecosystems have been altered since the onset of anthropogenic N deposition. 

15 As noted previously, studies found that N-limitation occurs as frequently as P-limitation in 

16 freshwater ecosystems (US EPA, 2008, section 3.3.3.2). Recently, a comprehensive study of 

17 available data from the northern hemisphere surveys of lakes along gradients of N deposition 

18 show increased inorganic N concentration and productivity to be correlated with atmospheric N 

19 deposition. The results are unequivocal evidence of N limitation in lakes with low ambient inputs 

20 of N, and increased N concentrations in lakes receiving N solely from atmospheric N deposition. 

21 It has been suggested that most lakes in the northern hemisphere may have originally been N22 limited, and that atmospheric N deposition has changed the balance of N and P in lakes.
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1 Available data suggest that the increases in total N deposition do not have to be large to 

2 elicit an ecological effect. For example, a hindcasting exercise determined that the change in 

3 Rocky Mountain National Park lake algae that occurred between 1850 and 1964 was associated 

4 with an increase in wet N deposition that was only about 1.5 kg N/ha. Similar changes inferred 

5 from lake sediment cores of the Beartooth Mountains of Wyoming also occurred at about 1.5 kg 

6 N/ha deposition. Pre-industrial inorganic N deposition is estimated to have been only 0.1 to 0.7 

7 kg N/ha based on measurements from remote parts of the world. In the western U.S., pre8 industrial, or background, inorganic N deposition was estimated by to range from 0.4 to 0.7 kg 

9 N/ha/yr.

10 Eutrophication effects from N deposition are most likely to be manifested in undisturbed, 

11 low nutrient surface waters such as those found in the higher elevation areas of the western U.S. 

12 The most severe eutrophication from N deposition effects is expected downwind of major urban 

and agricultural centers. High concentrations of lake or streamwater NO3

−

13 , indicative of 

14 ecosystem saturation, have been found at a variety of locations throughout the U.S., including the 

15 San Bernardino and San Gabriel Mountains within the Los Angeles Air Basin, the Front Range 

16 of Colorado, the Allegheny mountains of West Virginia, the Catskill Mountains of New York, 

17 the Adirondack Mountains of New York, and the Great Smoky Mountains in Tennessee (US 

18 EPA, 2008, section 3.3.8).

19 In contrast to terrestrial and freshwater systems, atmospheric N load to estuaries 

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

21 sources of N often greatly exceed N deposition. In estuaries, N-loading from multiple 

22 anthropogenic and non-anthropogenic pathways leads to water quality deterioration, resulting in 

23 numerous effects including hypoxic zones, species mortality, changes in community composition 
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1 and harmful algal blooms that are indicative of eutrophication. The following summary is a 

2 concise overview of the known or anticipated effects of nitrogen enrichment on estuaries within 

3 the United States.

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

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

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

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

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

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

10 rivers’ discharges to the Gulf of Mexico. 

11 A recent national assessment of eutrophic conditions in estuaries found that 65% of the 

12 assessed systems had moderate to high overall eutrophic conditions and generally received the 

13 greatest N loads from all sources, including atmospheric and land-based sources. Most eutrophic 

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

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

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

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

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

19 Submerged aquatic vegetation is important to the quality of estuarine ecosystem habitats because 

20 it provides habitat for a variety of aquatic organisms, absorbs excess nutrients, and traps 

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

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

23 negative impacts from nitrogen addition from atmospheric deposition.
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1 ii. Terrestrial ecosystems

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

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

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

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

6 likely to occur where areas of relatively high atmospheric N deposition intersect with N-limited 

7 plant communities. The alpine ecosystems of the Colorado Front Range, chaparral watersheds of 

8 the Sierra Nevada, lichen and vascular plant communities in the San Bernardino Mountains and 

9 the Pacific Northwest, and the southern California coastal sage scrub (CSS) community are 

10 among the most sensitive terrestrial ecosystems. There is growing evidence (US EPA, 2008, 

11 section 4.3.1.2) that existing grassland ecosystems in the western United States are being altered 

12 by elevated levels of N inputs, including inputs from atmospheric deposition.

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

assessed in terms of the degree of NO3

−

14 leaching from watershed soils into ground water or 

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

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

17 60% were in Stage 1 or Stage 2 of N saturation on a scale of 0 (background or pretreatment) to 3 

18 (visible decline). Of the northeastern sites for which adequate data were available for assessment, 

19 those in Stage 1 or 2 were most prevalent in the Adirondack and Catskill Mountains. Effects on 

20 individual plant species have not been well studied in the U.S. More is known about the 

21 sensitivity of particular plant communities. Based largely on results obtained in more extensive 

22 studies conducted in Europe, it is expected that the more sensitive terrestrial ecosystems include 
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1 hardwood forests, alpine meadows, arid and semi-arid lands, and grassland ecosystems (US 

2 EPA, 2008, section 3.3.5).

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

4 2008, Table 4.4) to identify meaningful ecological benchmarks associated with different levels of 

5 atmospheric nitrogen deposition. These are illustrated in Figure 3-4 of the Policy Assessment. 

6 The sensitive areas and ecological indicators identified by the ISA were analyzed further in the 

7 REA to create a national map that illustrates effects observed from ambient and experimental 

8 atmospheric nitrogen deposition loads in relation to Community Multi-scale Air Quality 

9 (CMAQ) 2002 modeling results and NADP monitoring data. This map, reproduced in Figure 3-

10 5 of the Policy Assessment, depicts the sites where empirical effects of terrestrial nutrient 

11 enrichment have been observed and site proximity to elevated atmospheric N deposition. 

12 Based on information in the ISA and initial analysis in the REA, further case study

13 analyses on terrestrial nutrient enrichment of ecosystems were developed for the CCS 

14 community and Mixed Conifer Forest (MCF) (US EPA, 2009). Geographic information systems 

15 (GIS) analysis supported a qualitative review of past field research to identify ecological 

16 benchmarks associated with CSS and mycorrhizal communities, as well as MCF nutrient17 sensitive acidophyte lichen communities, fine-root biomass in Ponderosa pine, and leached 

18 nitrate in receiving waters. 

19 The ecological benchmarks that were identified for the CSS and the MCF communities 

20 are included in the suite of benchmarks identified in the ISA (US EPA, 2008, section 3.3). There 

21 are sufficient data to confidently relate the ecological effect to a loading of atmospheric nitrogen. 

22 For the CSS community, the following ecological benchmarks were identified:
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1 (1) 3.3 kg N/ha/yr – the amount of nitrogen uptake by a vigorous stand of CSS; above this level, 

2 nitrogen may no longer be limiting

3 (2) 10 kg N/ha/yr – mycorrhizal community changes

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

5 (1) 3.1 kg N/ha/yr – shift from sensitive to tolerant lichen species

6 (2) 5.2 kg N/ha/yr – dominance of the tolerant lichen species

7 (3) 10.2 kg N/ha/yr – loss of sensitive lichen species

8 (4) 17 kg N/ha/yr – leaching of nitrate into streams.

9 These benchmarks, ranging from 3.1 to 17 kg N/ha/yr, were compared to 2002 

10 CMAQ/NADP data to discern any associations between atmospheric deposition and changing 

11 communities. Evidence supports the finding that nitrogen alters CSS and MCF communities. 

12 Key findings include the following: 2002 CMAQ/NADP nitrogen deposition data show that the 

13 3.3 kg N/ha/yr benchmark has been exceeded in more than 93% of CSS areas (654,048 ha). 

14 These deposition levels are a driving force in the degradation of CSS communities. Although 

15 CSS decline has been observed in the absence of fire, the contributions of deposition and fire to 

16 the CSS decline require further research. CSS is fragmented into many small parcels, and the 

17 2002 CMAQ/NADP 12-km grid data are not fine enough to fully validate the relationship 

18 between CSS distribution, nitrogen deposition, and fire. 2002 CMAQ/NADP nitrogen deposition 

19 data exceeds the 3.1 kg N/ha/yr benchmark in more than 38% (1,099,133 ha) of MCF areas, and 

20 nitrate leaching has been observed in surface waters. Ozone effects confound nitrogen effects on 

21 MCF acidophyte lichen, and the interrelationship between fire and nitrogen cycling requires 

22 additional research.
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1 c. Key uncertainties associated with nutrient enrichment

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

3 ecological effects and ecological indicators. The criteria used in the REA to evaluate the degree 

4 of confidence in the data, modeling and ecological effect function are detailed in Chapter 7 of the 

5 REA. Below is a discussion of uncertainty relating aquatic and terrestrial ecosystems to nutrient 

6 enrichment effects.

7 i. Aquatic ecosystems 

8 The approach for assessing atmospheric contributions to total nitrogen loading in the 

9 REA was to consider the main-stem river to an estuary (including the estuary) rather than an 

10 entire estuary system or bay. The biological indicators used in the NOAA ASSETS EI required 

11 the evaluation of many national databases including the US Geological Survey National Water 

12 Quality Assessment (NAWQA) files, EPA’s STORage and RETrieval (STORET) database, 

13 NOAA’s Estuarine Drainage Areas data, and EPA’s water quality standards nutrient criteria for 

14 rivers and lakes (US EPA, 2009, Appendix 6 and Table 1.2.-1). Both the SPARROW modeling 

15 for nitrogen loads and assessment of estuary conditions under NOAA ASSETS EI, have been 

16 applied on a national scale. The REA concludes that the available data are medium quality with 

17 intermediate confidence about the use of these data and their values for extrapolating to a larger 

18 regional area (US EPA, 2009, section 7.3.1). Intermediate confidence is associated with the 

19 modeling approach using ASSETS EI and SPARROW. The REA states there is low confidence 

20 with the ecological effect function due to the results of the analysis which indicated that 

21 reductions in atmospheric deposition alone could not solve coastal eutrophication problems due 

22 to multiple non-atmospheric nitrogen inputs (US EPA, 2009, section 7.3.3).
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1 ii. Terrestrial ecosystems

2 Ecological thresholds are identified for CSS and MCF areas and these data are considered 

3 to be of high quality, however, the ability to extrapolate these data to larger regional areas is 

4 limited (US EPA, 2009, section 7.4.1). No quantitative modeling was conducted or ecological 

5 effect function developed for terrestrial nutrient enrichment reflecting the uncertainties 

6 associated with these depositional effects. 

7 4. Other Ecological Effects 

8 It is stated in the ISA (US EPA, 2008, section 3.4.1 and 4.5) that mercury is a highly 

9 neurotoxic contaminant that enters the food web as a methylated compound, methylmercury. 

10 Mercury is principally methylated by sulfur-reducing bacteria and can be taken up by

11 microorganisms, zooplankton and macroinvertebrates. The contaminant is concentrated in higher 

12 trophic levels, including fish eaten by humans. Experimental evidence has established that only 

13 inconsequential amounts of methylmercury can be produced in the absence of sulfate. Once 

14 methylmercury is present, other variables influence how much accumulates in fish, but elevated 

15 mercury levels in fish can only occur where substantial amounts of methylmercury are present. 

16 Current evidence indicates that in watersheds where mercury is present, increased oxides of 

17 sulfur deposition very likely results in additional production of methylmercury which leads to 

18 greater accumulation of MeHg concentrations in fish. With respect to sulfur deposition and 

19 mercury methylation, the final ISA determined: The evidence is sufficient to infer a causal 

20 relationship between sulfur deposition and increased mercury methylation in wetlands and 

21 aquatic environments. 

22 The production of meaningful amounts of methylmercury (MeHg) requires the presence 

of SO4

2-

and mercury, and where mercury is present, increased availability of SO4

2-

23 results in 
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1 increased production of MeHg. There is increasing evidence on the relationship between sulfur 

2 deposition and increased methylation of mercury in aquatic environments; this effect occurs only 

3 where other factors are present at levels within a range to allow methylation. The production of 

4 methylmercury requires the presence of sulfate and mercury, but the amount of methylmercury

5 produced varies with oxygen content, temperature, pH, and supply of labile organic carbon (US 

6 EPA, 2008, section 3.4). In watersheds where changes in sulfate deposition did not produce an 

7 effect, one or several of those interacting factors were not in the range required for meaningful 

8 methylation to occur (US EPA, 2008, section 3.4). Watersheds with conditions known to be 

9 conducive to mercury methylation can be found in the northeastern United States and 

10 southeastern Canada. 

11 While the relationship between sulfur and methylmercury production was concluded to 

12 be causal in the ISA, the REA concluded that there was insufficient evidence to quantify the 

13 relationship between sulfur and methylmercury. Therefore only a qualitative assessment was 

14 included in Chapter 6 of the REA. The Policy Assessment was then unable to make a 

15 determination as to the adequacy of the existing SO2 standards in protecting against welfare 

16 effects associated with increased mercury methylation.

17 B. Risk and Exposure Assessment 

18 The risk and exposure assessment conducted for the current review was developed to 

19 describe potential risk from current and future deposition of oxides of nitrogen and sulfur to 

20 sensitive ecosystems. The case study analyses in the REA show that there is confidence that 

21 known or anticipated adverse ecological effects are occurring under current ambient loadings of 

22 nitrogen and sulfur in sensitive ecosystems across the United States. An overview of the 

23 material covered in the REA, a summary of the key findings from the air quality analyses, 
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1 acidification and nutrient enrichment case studies, and general conclusions from evaluating 

2 additional welfare effects, are presented below.

3 1. Overview of the Risk and Exposure Assessment

4 The REA evaluates the relationships between atmospheric concentrations, deposition, 

5 biologically relevant exposures, targeted ecosystem effects, and ecosystem services. To evaluate 

6 the nature and magnitude of adverse effects associated with deposition, the REA also examines 

7 various ways to quantify the relationships between air quality indicators, deposition of 

8 biologically available forms of nitrogen and sulfur, ecologically relevant indicators relating to 

9 deposition, exposure and effects on sensitive receptors, and related effects resulting in changes in 

10 ecosystem structure and services. The intent is to determine the exposure metrics that incorporate 

11 the temporal considerations (i.e., biologically relevant timescales), pathways, and ecologically 

12 relevant indicators necessary to determine the effects on these ecosystems. To the extent feasible, 

13 the REA evaluates the overall load to the system for nitrogen and sulfur, as well as the variability

14 in ecosystem responses to these pollutants. It also evaluates the contributions of atmospherically 

15 deposited nitrogen and sulfur individually relative to the combined atmospheric loadings of both 

16 elements together. Since oxidized nitrogen is the listed criteria pollutant (currently measured by 

17 the ambient air quality indicator NO2) for the atmospheric contribution to total nitrogen, the REA 

18 examines the contribution of nitrogen oxides to total reactive nitrogen in the atmosphere, relative 

19 to the contributions of reduced forms of nitrogen (e.g., ammonia, ammonium), to ultimately

20 assess how a meaningful secondary National Ambient Air Quality Standards (NAAQS) might be 

21 structured.

22 The REA focuses on ecosystem welfare effects that result from the deposition of total 

23 reactive nitrogen and sulfur. Because ecosystems are diverse in biota, climate, geochemistry, and 
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1 hydrology, response to pollutant exposures can vary greatly between ecosystems. In addition, 

2 these diverse ecosystems are not distributed evenly across the United States. To target nitrogen 

3 and sulfur acidification and nitrogen and sulfur enrichment, the REA addresses four main 

4 targeted ecosystem effects on terrestrial and aquatic systems identified by the ISA (US EPA, 

5 2008): aquatic acidification due to nitrogen and sulfur; terrestrial acidification due to nitrogen 

6 and sulfur; aquatic nutrient enrichment, including eutrophication; and terrestrial nutrient 

7 enrichment.

8 In addition to these four targeted ecosystem effects, the REA also qualitatively addresses 

9 the influence of sulfur oxides deposition on methylmercury production; nitrous oxide (N2O) 

10 effects on climate; nitrogen effects on primary productivity and biogenic greenhouse gas fluxes; 

11 and phytotoxic effects on plants. 

12 Because the targeted ecosystem effects outlined above are not evenly distributed across 

13 the United States, the REA identified case studies for each targeted effects based on ecosystems 

14 identified as sensitive to nitrogen and/or sulfur deposition effects. Eight case study areas and 

15 two supplemental study areas (Rocky Mountain National Park and Little Rock Lake, WI) are 

16 summarized in the REA based on ecosystem characteristics, indicators, and ecosystem service 

17 information. Case studies selected for aquatic acidification effects were the Adirondack 

18 Mountains and Shenandoah National Park. Kane Experimental Forest in Pennsylvania and 

19 Hubbard Brook Experimental Forest in New Hampshire were selected as case studies for 

20 terrestrial acidification. Aquatic nutrient enrichment case study locations were selected in the 

21 Potomac River Basin upstream of Chesapeake Bay and the Neuse River Basin upstream of the 

22 Pamlico Sound in North Carolina. The Coastal Sage Scrub Communities in southern California 

23 and the Mixed Confer Forest Communities in the San Bernardino and Sierra Nevada Mountains 
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1 of California were selected as case studies for terrestrial nutrient enrichment. Two supplemental 

2 areas were also chosen, one in Rocky Mountain National Park for terrestrial nutrient enrichment 

3 and one in Little Rock Lake, Wisconsin for aquatic nutrient enrichment.

4 2. Key findings

5 In summary, based on case study analyses, the REA concludes that known or anticipated 

6 adverse ecological effects are occurring under current conditions and further concludes that these 

7 adverse effects continue into the future. Key findings from the air quality analyses, acidification 

8 and nutrient enrichment case studies, as well as general conclusions from evaluating additional 

9 welfare effects, are summarized below.

10 a. Air quality analyses

11 The air quality analyses in the REA encompass the current emissions sources of nitrogen 

12 and sulfur, as well as atmospheric concentrations, estimates of deposition of total nitrogen, 

13 policy-relevant background, and non-ambient loadings of nitrogen and sulfur to ecosystems, both 

14 nationwide and in the case study areas. Spatial fields of deposition were created using wet 

15 deposition measurements from the National Atmospheric Deposition Program (NADP) National 

16 Trends Network and dry deposition predictions from the 2002 Community Multi-Scale Air 

17 Quality (CMAQ) model simulation. Some key conclusions from this analysis are:

18 (1) Total reactive nitrogen deposition and sulfur deposition are much greater in the East 

19 compared to most areas of the West.

20 (2) These regional differences in deposition correspond to the regional differences in oxides of 

21 nitrogen and SO2 concentrations and emissions, which are also higher in the East. Oxides of 

22 nitrogen emissions are much greater and generally more widespread than NH3 emissions 

23 nationwide; high NH3 emissions tend to be more local (e.g., eastern North Carolina) or sub-
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1 regional (e.g., the upper Midwest and Plains states). The relative amounts of oxidized versus 

2 reduced nitrogen deposition are consistent with the relative amounts of oxides of nitrogen

3 and NH3 emissions. Oxidized nitrogen deposition exceeds reduced nitrogen deposition in 

4 most of the case study areas; the major exception being the Neuse River/Neuse River Estuary

5 Case Study Area.

6 (3) Reduced nitrogen deposition exceeds oxidized nitrogen deposition in the vicinity of local 

7 sources of NH3.

8 (4) There can be relatively large spatial variations in both total reactive nitrogen deposition and

9 sulfur deposition within a case study area; this occurs particularly in those areas that contain 

10 or are near a high emissions source of oxides of nitrogen, NH3, and/or SO2.

11 (5) The seasonal patterns in deposition differ between the case study areas. For the case study 

12 areas in the East, the season with the greatest amounts of total reactive nitrogen deposition 

13 correspond to the season with the greatest amounts of sulfur deposition. Deposition peaks in 

14 spring in the Adirondack, Hubbard Brook Experimental Forest, and Kane Experimental 

15 Forest case study areas, and it peaks in summer in the Potomac River/Potomac Estuary, 

16 Shenandoah, and Neuse River/Neuse River Estuary case study areas. For the case study areas 

17 in the West, there is less consistency in the seasons with greatest total reactive nitrogen and 

18 sulfur deposition in a given area. In general, both nitrogen and/or sulfur deposition peaks in 

19 spring or summer. The exception to this is the Sierra Nevada Range portion of the Mixed 

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

21 b. Deposition-related aquatic acidification

22 The role of aquatic acidification in two eastern United States areas—northeastern New 

23 York’s Adirondack area and the Shenandoah area in Virginia—was analyzed in the REA to 
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assess surface water trends in SO4

2-

and NO3

-

1 concentrations and acid neutralizing capacity (ANC) 

2 levels and to affirm the understanding that reductions in deposition could influence the risk of 

3 acidification. Monitoring data from the EPA-administered Temporally Integrated Monitoring of 

4 Ecosystems (TIME)/Long-Term Monitoring (LTM) programs and the Environmental Monitoring 

5 and Assessment Program (EMAP) were assessed for the years 1990 to 2006, and past, present, 

6 and future water quality levels were estimated using both steady-state and dynamic 

7 biogeochemical models. 

8 Although wet deposition rates for SO2 and oxides of nitrogen in the Adirondack Case 

9 Study Area have reduced since the mid-1990s, current concentrations are still well above pre10 acidification (1860) conditions. Model of Acidification of Groundwater in Catchments (MAGIC) 

modeling predicts NO3

-

and SO4

2-

11 are 17- and 5-fold higher today, respectively. The estimated 

12 average ANC for 44 lakes in the Adirondack Case Study Area is 62.1 microequivalents per liter 

13 (μeq/L) (± 15.7 μeq/L); 78 % of all monitored lakes in the Adirondack Case Study Area have a 

14 current risk of Elevated, Severe, or Acute. Of the 78%, 31% experience episodic acidification, 

15 and 18% are chronically acidic today.

16 (1) Based on the steady-state critical load model for the year 2002, 18%, 28%, 44%, and 58% of 

17 169 modeled lakes received combined total sulfur and nitrogen deposition that exceeded

18 critical loads corresponding to ANC limits of 0, 20, 50, and 100 μeq/L respectively.

19 (2) Based on a deposition scenario that maintains current emission levels to 2020 and 2050, the 

20 simulation forecast indicates no improvement in water quality in the Adirondack Case Study 

21 Area. The percentage of lakes within the Elevated to Acute Concern classes remains the same 

22 in 2020 and 2050.
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1 (3) Since the mid-1990s, streams in the Shenandoah Case Study Area have shown slight declines 

in NO3 and SO4

2-

2 concentrations in surface waters. ANC levels increased from about 50 

3 μeq/L in the early 1990s to >75 μeq/L until 2002, when ANC levels declined back to 1991–

4 1992 levels. Current concentrations are still above pre-acidification (1860) conditions. 

MAGIC modeling predicts surface water concentrations of NO3 and SO4

2-

5 are 10- and 32-

6 fold higher today, respectively. The estimated average ANC for 60 streams in the 

7 Shenandoah Case Study Area is 57.9 μeq/L (± 4.5 μeq/L). 55% of all monitored streams in 

8 the Shenandoah Case Study Area have a current risk of Elevated, Severe, or Acute. Of the 

9 55%, 18% experience episodic acidification, and 18% are chronically acidic today.

10 (4) Based on the steady-state critical load model for the year 2002, 52%, 72%, 85%, and 93% of 

11 60 modeled streams received combined total sulfur and nitrogen deposition that exceeded 

12 critical loads corresponding to ANC limits of 0, 20, 50, and 100 μeq/L respectively.

13 (5) Based on a deposition scenario that maintains current emission levels to 2020 and 2050, the 

14 simulation forecast indicates that a large number of streams still have Elevated to Acute 

15 problems with acidity. In fact, from 2006 to 2050, the percentage of streams with Acute 

16 Concern increases by 5%, while the percentage of streams in Moderate Concern decreases 

17 by 5%.

18 c. Deposition-related terrestrial acidification

19 The role of terrestrial acidification was examined in the REA using a critical load 

20 analysis for sugar maple and red spruce forests in the eastern United States by using the base 

21 cation to aluminum (Bc/Al) ratio in acidified forest soils as an indicator to assess the impact of 

22 nitrogen and sulfur deposition on tree health. These are the two most commonly studied species 

23 in North America for impacts of acidification. At a Bc/Al ratio of 1.2, red spruce growth can be 
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1 reduced by 20%. Sugar maple growth can be reduced by 20% at a Bc/Al ratio of 0.6. Key 

2 findings of the case study are summarized below.

3 (1) Case study results suggest that the health of at least a portion of the sugar maple and red 

4 spruce growing in the United States may have been compromised with acidifying total 

5 nitrogen and sulfur deposition in 2002. 2002 CMAQ/NADP total nitrogen and sulfur 

6 deposition levels exceeded three selected critical loads in 3% to 75% of all sugar maple plots 

7 across 24 states. The three critical loads ranged from 107 to 6,008 eq/ha/yr for the Bc/Al 

8 ratios of 0.6, 1.2, and 10.0 (increasing levels of tree protection). The 2002 CMAQ/NADP 

9 total nitrogen and sulfur deposition levels exceeded three selected critical loads in 3% to 36% 

10 of all red spruce plots across 8 states. The three critical loads ranged from 180 to 4,278 

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

12 (2) The Simple Mass Balance model assumptions made for base cation weathering (Bcw) and 

13 forest soil ANC input parameters are the main sources of uncertainty since these parameters 

14 are rarely measured and require researchers to use default values. Bcw contributed 49% to 

15 the total variability in the critical load estimates, and forest soil ANC contributed 46% to the 

16 total variability.

17 (3) The pattern of case study results suggests that nitrogen and sulfur acidifying deposition in the 

18 sugar maple and red spruce forest areas studied were very close to, if not greater than, the 

19 critical loads for those areas and both ecosystems are likely to be sensitive to any future 

20 changes in the levels of deposition.

21 d. Deposition-related aquatic nutrient enrichment

22 The role of nitrogen deposition in two main stem rivers feeding their respective estuaries 

23 was analyzed in the REA to determine if decreases in deposition could influence the risk of 
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1 eutrophication as predicted using the Assessment of Estuarine Trophic Status eutrophication 

2 index (ASSETS EI) scoring system in tandem with SPAtially Referenced Regression on 

3 Watershed Attributes (SPARROW) modeling. This modeling approach provides a transferrable, 

4 intermediate-level analysis of the linkages between atmospheric deposition and receiving waters, 

5 while providing results on which conclusions could be drawn. Future application of the methods 

6 to case study areas where atmospheric deposition plays a larger role in the nitrogen loading to an 

7 estuary will likely provide more tangible results. A summary of findings follows:

8 (1) 2002 CMAQ/NADP results showed that an estimated 40,770,000 kg of total nitrogen was 

9 deposited in the Potomac River watershed. SPARROW modeling predicted that 7,380,000 kg 

10 N/yr of the deposited nitrogen reached the estuary (20% of the total load to the estuary). The 

11 overall ASSETS EI for the Potomac River and Potomac Estuary was Bad (based on all 

12 sources of N). 

13 (2) To improve the Potomac River and Potomac Estuary ASSETS EI score from Bad to Poor, 

14 there is a slim chance that a decrease of at least 78% in the 2002 total nitrogen atmospheric 

15 deposition load to the watershed would be required.

16 (3) 2002 CMAQ/NADP results showed that an estimated 18,340,000 kg of total nitrogen was 

17 deposited in the Neuse River watershed. SPARROW modeling predicted that 1,150,000 kg 

18 N/yr of the deposited nitrogen reached the estuary (26% of the total load to the estuary). The 

19 overall ASSETS EI for the Neuse River/Neuse River Estuary was Bad.

20 (4) It was found that the Neuse River/Neuse River Estuary ASSETS EI score could not be 

21 improved from Bad to Poor with decreases only in the 2002 atmospheric deposition load to 

22 the watershed. Additional reductions would be required from other nitrogen sources within 

23 the watershed.
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1 The small effect of decreasing atmospheric deposition in the Neuse River watershed is 

2 because the other nitrogen sources within the watershed are more influential than atmospheric 

3 deposition to the total nitrogen loadings to the Neuse River Estuary as estimated with the 

4 SPARROW model. A waterbody’s response to nutrient loading depends on the magnitude (e.g., 

5 agricultural sources have a high influence in the Neuse), spatial distribution, and other 

6 characteristics of the sources within the watershed.

7 e. Deposition-related terrestrial nutrient enrichment

8 California Coastal Sage Scrub (CSS) and Mixed Conifer Forest (MCF) communities 

9 were the focus of the Terrestrial Nutrient Enrichment Case Studies of the REA. Geographic 

10 information systems (GIS) analysis supported a qualitative review of past field research to 

11 identify ecological benchmarks associated with CSS and mycorrhizal communities, as well as 

12 MCF’s nutrient-sensitive acidophyte lichen communities, fine-root biomass in Ponderosa pine, 

13 and leached nitrate in receiving waters. These benchmarks, ranging from 3.1 to 17 kg N/ha/yr, 

14 were compared to 2002 CMAQ/NADP data to discern any associations between atmospheric 

15 deposition and changing communities. Evidence supports the finding that nitrogen alters CSS 

16 and MCF. Key findings include the following:

17 (1) 2002 CMAQ/NADP nitrogen deposition data show that the 3.3 kg N/ha/yr benchmark has 

18 been exceeded in more than 93% of CSS areas (654,048 ha). These deposition levels are a 

19 driving force in the degradation of CSS communities. Although CSS decline has been 

20 observed in the absence of fire, the contributions of deposition and fire to the CSS decline 

21 require further research. CSS is fragmented into many small parcels, and the 2002 

22 CMAQ/NADP 12-km grid data are not fine enough to fully validate the relationship between 

23 CSS distribution, nitrogen deposition, and fire.
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1 (2) 2002 CMAQ/NADP nitrogen deposition data exceeds the 3.1 kg N/ha/yr benchmark in more 

2 than 38% (1,099,133 ha) of MCF areas, and nitrate leaching has been observed in surface 

3 waters. Ozone effects confound nitrogen effects on MCF acidophyte lichen, and the 

4 interrelationship between fire and nitrogen cycling requires additional research.

5 f. Additional effects

6 Ecological effects have also been documented across the United States where elevated 

7 nitrogen deposition has been observed, including the eastern slope of the Rocky Mountains 

8 where shifts in dominant algal species in alpine lakes have occurred where wet nitrogen 

9 deposition was only about 1.5 kg N/ha/yr. High alpine terrestrial communities have a low 

10 capacity to sequester nitrogen deposition, and monitored deposition exceeding 3 to 4 kg N/ha/yr 

11 could lead to community-level changes in plant species, lichens, and mycorrhizae.

12 Additional welfare effects are documented, but examined less extensively, in the REA. 

13 These effects include qualitative discussions related to visibility and materials damage, such as 

14 corrosion, erosion, and soiling of paint and buildings which are being addressed in the particulate 

15 matter (PM) NAAQS review currently underway. A discussion of the causal relationship 

between sulfur deposition (as sulfate, SO4

2-

16 ) and increased mercury methylation in wetlands and 

17 aquatic environments is also included in the REA. On this subject the REA concludes that 

18 decreases in sulfate deposition will likely result in decreases in methyl mercury concentration; 

19 however, spatial and biogeochemical variations nationally hinder establishing large scale dose20 response relationships. 

21 Several additional issues concerning oxides of nitrogen were addressed in the REA. 

22 Consideration was also given to nitrous oxide (N2O), a potent GHG. The REA concluded that it 

23 is most appropriate to analyze the role of N2O in the context of all of the GHGs rather than as 
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1 part of the REA for this review. The REA considered nitrogen deposition and its correlation 

2 with the rate of photosynthesis and net primary productivity. Nitrogen addition ranging from 

3 15.4 to 300 kg N/ha/yr is documented as increasing wetland N2O production by an average of 

4 207%. Nitrogen addition ranging from 30 to 240 kg N/ha/y increased CH4 emissions by 115%, 

5 averaged across all ecosystems, and methane uptake was reduced by 38% when nitrogen addition 

6 ranged from 10 to 560 kg N/ha/yr, but reductions were only significant for coniferous and 

7 deciduous forests. The heterogeneity of ecosystems across the United States, however, 

8 introduces variations into dose-response relationships.

9 The phytotoxic effects of oxides of nitrogen and sulfur on vegetation were also briefly 

10 discussed in the REA which concluded that since a unique secondary NAAQS exists for SO2, 

11 and concentrations of nitric oxide (NO), NO2, and PAN are rarely high enough to have 

12 phytotoxic effects on vegetation, further assessment was not warranted at this time. 

13 3. Conclusions on Effects 

14 For aquatic and terrestrial acidification effects, a similar conceptual approach was used 

15 (critical loads) to evaluate the impacts of multiple pollutants on an ecological endpoint, whereas 

16 the approaches used for aquatic and terrestrial nutrient enrichment were fundamentally distinct. 

17 Although the ecological indicators for aquatic and terrestrial acidification (i.e., ANC and Bc/Al) 

18 are very different, both ecological indicators are well-correlated with effects such as reduced 

19 biodiversity and growth. While aquatic acidification is clearly the targeted effect area with the 

20 highest level of confidence, the relationship between atmospheric deposition and an ecological 

21 indicator is also quite strong for terrestrial acidification. The main drawback with the 

22 understanding of terrestrial acidification is that the data are based on laboratory responses rather 
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1 than field measurements. Other stressors that are present in the field but that are not present in 

2 the laboratory may confound this relationship.

3 For nutrient enrichment effects, the REA utilized different types of indicators for aquatic 

4 and terrestrial effects to assess both the likelihood of adverse effects to ecosystems and the 

5 relationship between adverse effects and atmospheric sources of oxides of nitrogen. The 

6 ecological indicator chosen for aquatic nutrient enrichment, the ASSETS EI, seems to be 

7 inadequate to relate atmospheric deposition to the targeted ecological effect, likely due to the 

8 many other confounding factors. Further, there is far less confidence associated with the 

9 understanding of aquatic nutrient enrichment because of the large contributions from non10 atmospheric sources of nitrogen and the influence of both oxidized and reduced forms of 

11 nitrogen, particularly in large watersheds and coastal areas. However, a strong relationship exists 

12 between atmospheric deposition of nitrogen and ecological effects in high alpine lakes in the 

13 Rocky Mountains because atmospheric deposition is the only source of nitrogen to these 

14 systems. There is also a strong weight-of-evidence regarding the relationships between 

15 ecological effects attributable to terrestrial nitrogen nutrient enrichment; however, ozone and 

16 climate change may be confounding factors. In addition, the response for other species or species 

17 in other regions of the United States has not been quantified.

18 C. Adversity of Effects to Public Welfare

19 Characterizing a known or anticipated adverse effect to public welfare is an important 

20 component of developing any secondary NAAQS. According to the Clean Air Act, welfare 

21 effects include: “effects on soils, water, crops, vegetation, manmade materials, animals, wildlife, 

22 weather, visibility, and climate, damage to and deterioration of property, and hazards to 

23 transportation, as well as effect on economic values and on personal comfort and well-being, 
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1 whether caused by transformation, conversion, or combination with other air pollutants” (CAA, 

2 Section 302(h)). While the text above lists a number of welfare effects, these effects do not 

3 define public welfare in and of themselves. 

4 Although there is no specific definition of adversity to public welfare, the paradigm of 

5 linking adversity to public welfare to disruptions in ecosystem structure and function has been 

6 used broadly by EPA to categorize effects of pollutants from the cellular to the ecosystem level. 

7 An evaluation of adversity to public welfare might consider the likelihood, type, magnitude, and 

8 spatial scale of the effect as well as the potential for recovery and any uncertainties relating to 

9 these considerations. 

10 Similar concepts were used in past reviews of secondary NAAQS for ozone and PM 

11 (relating to visibility), as well as in initial reviews of effects from lead deposition. Because 

12 oxides of nitrogen and sulfur are deposited from ambient sources into ecosystems where they

13 affect changes to organisms, populations and ecosystems, the concept of adversity to public 

14 welfare as a result of alterations in structure and function of ecosystems is an appropriate 

15 consideration for this review. 

16 Based on information provided in the Policy Assessment, the following section discusses 

17 how ecological effects from deposition of oxides of nitrogen and sulfur relate to adversity to 

18 public welfare. In the Policy Assessment, public welfare was discussed in terms of loss of 

19 ecosystem services (defined below), which in some cases can be monetized. Each of the four 

20 main effect areas (aquatic and terrestrial acidification and aquatic and terrestrial nutrient over21 enrichment) are discussed including current ecological effects and associated ecosystem services, 

22 including potential value of the services. The quantified indicators of value discussed below 
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1 relate to ecosystem services generally that are impacted by deposition related acidification, not to 

2 the specific effects of this action.

3 1. Ecosystem Services 

4 The Policy Assessment defines ecosystem services as the benefits individuals and 

5 organizations obtain from ecosystems. Ecosystem services can be classified as provisioning 

6 (food and water), regulating (control of climate and disease), cultural (recreational, existence, 

7 spiritual, educational), and supporting (nutrient cycling). Conceptually, changes in ecosystem 

8 services may be used to aid in characterizing a known or anticipated adverse effect to public 

9 welfare. In the REA and Policy Assessment ecosystem services are discussed as a method of 

10 assessing the magnitude and significance to the public of resources affected by ambient 

11 concentrations of oxides of nitrogen and sulfur and deposition in sensitive ecosystems. 

12 EPA has in previous NAAQS reviews defined ecological goods and services for the 

13 purposes of a Regulatory Impact Analysis as the “outputs of ecological functions or processes 

14 that directly or indirectly contribute to social welfare or have the potential to do so in the future. 

15 Some outputs may be bought and sold, but most are not marketed.” It is especially important to

16 acknowledge that it is difficult to measure and/or monetize the goods and services supplied by

17 ecosystems. It can be informative in characterizing adversity to public welfare to attempt to 

18 place an economic valuation on the set of goods and services that have been identified with 

19 respect to a change in policy however it must be noted that this valuation will be incomplete and 

20 illustrative only. 

21 Knowledge about the relationships linking ambient concentrations and ecosystem 

22 services is considered in the Policy Assessment as one method by which to inform a policy

23 judgment on a known or anticipated adverse public welfare effect. For example, a change in an 
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1 ecosystem structure and process, such as foliar injury, would be classified as an ecological effect, 

2 with the associated changes in ecosystem services, such as primary productivity, food 

3 availability, forest products, and aesthetics (e.g., scenic viewing), classified as public welfare 

4 effects. Additionally, changes in biodiversity would be classified as an ecological effect, and the 

5 associated changes in ecosystem services—productivity, existence (nonuse) value, recreational 

6 viewing and aesthetics—would also be classified as public welfare effects. 

7 As described in Chapters 4 and 5 of the REA, case study analyses were performed that 

8 link deposition in sensitive ecosystems to changes in a given ecological indicator (e.g., for 

9 aquatic acidification, to changes in acid neutralizing capacity [ANC]) and then to changes in 

10 ecosystems. Appendix 8 of the REA links the changes in ecosystems to the services they

11 provide (e.g., fish species richness and its influence on recreational fishing). To the extent 

12 possible for each targeted effect area, the REA linked ambient concentrations of nitrogen and 

13 sulfur (i.e., ambient air quality indicators) to deposition in sensitive ecosystems (i.e., exposure 

14 pathways), and then to system response as measured by a given ecological indicator (e.g., lake 

15 and stream acidification as measured by ANC). The ecological effect (e.g., changes in fish 

16 species richness) was then, where possible, associated with changes in ecosystem services and 

17 the corresponding public welfare effects (e.g., recreational fishing). 

18 2. Effects on Ecosystem Services

19 The process used to link ecological indicators to ecosystem services is discussed 

20 extensively in Appendix 8 of the REA. In brief, for each case study area assessed the ecological 

21 indicators are linked to an ecological response that is subsequently linked to associated services 

22 to the extent possible. For example in the case study for aquatic acidification the chosen 

23 ecological indicator is ANC which can be linked to the ecosystem service of recreational fishing. 
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1 Although recreational fishing losses are the only service effects that can be independently 

2 quantified or monetized at this time, there are numerous other ecosystem services that may be 

3 related to the ecological effects of acidification.

4 While aquatic acidification is the focus of this proposed standard, the other effect areas 

5 were also analyzed in the REA and these ecosystems are being harmed by nitrogen and sulfur 

6 deposition and will obtain some measure of protection with any decrease in that deposition 

7 regardless of the reason for the decrease. The following summarizes the current levels of specific 

8 ecosystem services for aquatic and terrestrial acidification, and aquatic and terrestrial nutrient 

9 over-enrichment and attempts to quantify and when possible monetize the harm to public 

10 welfare, as represented by ecosystem services, due to nitrogen and sulfur deposition. 

11 a. Aquatic Acidification

12 Acidification of aquatic ecosystems primarily affects the ecosystem services that are 

13 derived from the fish and other aquatic life found in surface waters. In the northeastern United 

14 States, the surface waters affected by acidification are not a major source of commercially raised 

15 or caught fish; however, they are a source of food for some recreational and subsistence fishers 

16 and for other consumers. Although data and models are available for examining the effects on 

17 recreational fishing, relatively little data are available for measuring the effects on subsistence 

18 and other consumers. Inland waters also provide aesthetic and educational services along with 

19 non-use services, such as existence value (protection and preservation with no expectation of 

20 direct use). In general, inland surface waters such as lakes, rivers, and streams also provide a 

21 number of regulating services, playing a role in hydrological regimes and climate regulation. 

22 There is little evidence that acidification of freshwaters in the northeastern United States has 

23 significantly degraded these specific services; however, freshwater ecosystems also provide 
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1 biological control services by providing environments that sustain delicate aquatic food chains. 

2 The toxic effects of acidification on fish and other aquatic life impair these services by disrupting 

3 the trophic structure of surface waters. Although it is difficult to quantify these services and how 

4 they are affected by acidification, it is worth noting that some of these services may be captured 

5 through measures of provisioning and cultural services. For example, these biological control 

6 services may serve as “intermediate” inputs that support the production of “final” recreational 

7 fishing and other cultural services.

8 As summarized in Chapter 4 of the Policy Assessment, recent studies indicate that 

9 acidification of lakes and streams can result in significant loss in economic value. For example, 

10 data indicate that more than 9% of adults in the northeastern part of the country participate 

11 annually in freshwater fishing yielding 140 million freshwater fishing days. Each fishing day

12 has an estimated average value per day of $35. Therefore, the implied total annual value of 

13 freshwater fishing in the northeastern United States was $5 billion in 2006. Embedded in these 

14 numbers is a degree of harm to recreational fishing services due to acidification that has occurred 

15 over time. These harms have not been quantified on a regional scale; however, a case study was 

16 conducted in the Adirondacks area (US EPA, 2011, section 4.4.2). 

17 In the Adirondacks case study, estimates of changes in recreational fishing services were 

18 determined, as well as changes more broadly in “cultural” ecosystem services (including 

19 recreational, aesthetic, and nonuse services). First, the MAGIC model (US EPA, 2009,

20 Appendix 8 and section 2.2) was applied to 44 lakes to predict what ANC levels would be under 

21 both “business as usual” conditions (i.e., allowing for some decline in deposition due to existing 

22 regulations) and pre-emission (i.e., background) conditions. Second, to estimate the recreational 

23 fishing impacts of aquatic acidification in these lakes, an existing model of recreational fishing 
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1 demand and site choice was applied. This model predicts how recreational fishing patterns in the 

2 Adirondacks would differ and how much higher the average annual value of recreational fishing 

3 services would be for New York residents if lake ANC levels corresponded to background 

4 (rather than business as usual) conditions. To estimate impacts on a broader category of cultural 

5 (and some provisioning) ecosystem services, results from the Banzhaf et al (2006) valuation 

6 survey of New York residents were adapted and applied to this context. The survey used a 

7 contingent valuation approach to estimate the average annual household WTP for future 

8 reductions in the percent of Adirondack lakes impaired by acidification. The focus of the survey 

9 was on impacts on aquatic resources. Pretesting of the survey indicated that respondents 

10 nonetheless tended to assume that benefits would occur in the condition of birds and forests as 

11 well as in recreational fishing. 

12 By extrapolating the 44 lake Adiriondack case study to all 3,000 Adirondack lakes and by

13 applying the willingness to pay survey results to all New York residents, the study estimated 

14 aggregated benefits between $300 and $800 million annually for the equivalent of improving 

15 lakes in the Adirondacks region to an ANC level of 50 μeq/L. The REA estimated 44% of the 

16 Adirondack lakes currently fall below an ANC of 50 μeq/L. Several states have set goals for 

17 improving the acid status of lakes and streams, generally targeting ANC in the range of 50 to 60 

18 μeq/L, and have engaged in costly activities to decrease acidification. 

19 These results imply significant value to the public in addition to those derived from 

20 recreational fishing services. Note that the results are only applicable to improvements in the 

21 Adirondacks valued by residents of New York. If similar benefits exist in other acid-impacted 

22 areas, benefits for the nation as a whole could be substantial. The analysis provides results on 
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1 only a subset of the impacts of acidification on ecosystem services and suggests that the overall 

2 impact on these services is likely to be substantial.

3 b. Terrestrial Acidification

4 Chapter 4.4.3 and 4.4.4 of the Policy Assessment review several economic studies of 

5 areas sensitive to terrestrial acidification. Forests in the northeastern United States provide 

6 several important and valuable provisioning ecosystem services, which are reflected in the 

7 production and sales of tree products. Sugar maples are a particularly important commercial 

8 hardwood tree species in the United States, producing timber and maple syrup that provide 

9 hundreds of millions of dollars in economic value annually. Red spruce is also used in a variety 

10 of wood products and provides up to $100 million in economic value annually. Although the 

11 data do not exist to directly link acidification damages to economic values of lost recreational 

12 ecosystem services in forests, these resources are valuable to the public. A recent study, 

13 reviewed in the Policy Assessment, suggests that the total annual value of recreational off-road 

14 driving was more than $9 billion and the value of hunting and wildlife viewing was more than $4 

15 billion each in the northeastern United States. EPA is not able to quantify at this time the 

16 specific effects on these values of acid deposition, or of any specific reductions in deposition,

17 relative to the effects of many other factors that may affect them.

18 c. Nutrient Enrichment

19 Chapter 4.4.5 and 4.4.6 of the Policy Assessment summarize economic studies of east 

20 coast estuaries affected by nutrient over-enrichment or eutrophication. Estuaries in the eastern 

21 United States are important for fish and shellfish production. The estuaries are capable of 

22 supporting large stocks of resident commercial species, and they serve as the breeding grounds 

23 and interim habitat for several migratory species. To provide an indication of the magnitude of 
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1 provisioning services associated with coastal fisheries, from 2005 to 2007, the average value of 

2 total catch was $1.5 billion per year in 15 East Coast states. Estuaries also provide an important 

3 and substantial variety of cultural ecosystem services, including water-based recreational and 

4 aesthetic services. For example, data indicate that 4.8% of the population in coastal states from 

5 North Carolina to Massachusetts participated in saltwater fishing, with a total of 26 million 

6 saltwater fishing days in 2006. Based on estimates in the Policy Assessment, total recreational 

7 value from these saltwater fishing days was approximately $1.3 billion. Recreational 

8 participation estimates for 1999–2000 showed almost 6 million individuals participated in 

9 motorboating in coastal states from North Carolina to Massachusetts. The aggregate value of 

10 these coastal motorboating outings was $2 billion per year. EPA is not able to quantify at this 

11 time the specific effects on these values of acid deposition, or of any specific reductions in 

12 deposition, relative to the effects of many other factors that may affect them.

13 Terrestrial ecosystems can also suffer from nutrient over-enrichment. Each ecosystem is 

14 different in its composition of species and nutrient requirements. Changes to individual 

15 ecosystems from changes in nitrogen deposition can be hard to assess economically. Relative 

16 recreational values are often determined by public use information. Chapter 4.4.7 of the Policy

17 Assessment reviewed studies related to park use in California. Data from California State Parks 

18 indicate that in 2002, 68.7% of adult residents participated in trail hiking for an average of 24.1 

19 days per year. The analyses in the Policy Assessment indicate that the aggregate annual benefit 

20 for California residents from trail hiking in 2007 was $11.59 billion. EPA is not able to quantify 

21 at this time the specific effects on these values of acid deposition, or of any specific reductions in 

22 deposition, relative to the effects of many other factors that may affect them.
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1 The Policy Assessment also identified fire regulation as a service that could be affected 

2 by nutrient over-enrichment of the Coastal Sage Scrub and Mixed Conifer Forest ecosystems by

3 encouraging growth of more flammable grasses, increasing fuel loads, and altering the fire cycle. 

4 Over the 5-year period from 2004 to 2008, Southern California experienced, on average, over 

5 4,000 fires per year, burning, on average, over 400,000 acres per year. It is not possible at this 

6 time to quantify the contribution of nitrogen deposition, among many other factors, to increased 

7 fire risk.

8 3. Summary

9 Adversity to public welfare can be understood by looking at how deposition of oxides of 

10 nitrogen and sulfur affect the ecological functions of an ecosystem (see II.A.), and then

11 understanding the ecosystem services that are degraded. The monetized value of the ecosystem 

12 services provided by ecosystems that are sensitive to deposition of oxides of nitrogen and sulfur 

13 are in the billions of dollars each year, though it is not possible to quantify or monetize at this 

14 time the effects on these values of acid deposition or of any changes in deposition that may result 

15 from new secondary standards. Many lakes and streams are known to be degraded by acidic 

16 deposition which affects recreational fishing and tourism. Forest growth is likely suffering from 

17 acidic deposition in sensitive areas affecting red spruce and sugar maple timber production, sugar 

18 maple syrup production, hiking, aesthetic enjoyment and tourism. Nitrogen deposition 

19 contributes significantly to eutrophication in many estuaries affecting fish production, 

20 swimming, boating, aesthetic enjoyment, and tourism. Ecosystem services are likely affected by

21 nutrient enrichment in many natural and scenic terrestrial areas, affecting biodiversity, including 

22 habitat for rare and endangered species, fire control, hiking, aesthetic enjoyment, and tourism. 

23 D. Adequacy of the Current Standards
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1 An important issue to be addressed in the current review of the secondary standards for 

2 oxides and nitrogen and sulfur standard is whether, in view of the scientific evidence reflected in 

3 the ISA, additional information on exposure and risk discussed in the REA, and conclusions 

4 drawn from the Policy Assessment, the existing standards provide adequate protection. The 

5 Administrator therefore, has considered the extent to which the current standards are adequate 

6 for the protection of public welfare. Having reached the general conclusion that aquatic and 

7 terrestrial ecosystems can be degrade by deposition of oxides of nitrogen and sulfur, it is then 

8 necessary to evaluate the adequacy of the current secondary standards for oxides of nitrogen and 

9 sulfur to provide requisite protection by considering to what degree risks to sensitive ecosystems 

10 would be expected to occur in areas that meet the current standards. Conclusions regarding the 

11 adequacy of the current standards are based on the available ecological effects, exposure and 

12 risk-based evidence. In evaluating the strength of this information, EPA has taken into account 

13 the uncertainties and limitations in the scientific evidence. This section addresses the adequacy

14 of the current standards to protect against direct exposure effects on plants from oxides of 

15 nitrogen and sulfur, the appropriateness of the current structure of the standards to address 

16 deposition-related effects of oxides of nitrogen and sulfur on sensitive ecosystems and finally, 

17 the adequacy of such standards to protect against adverse effects related to the deposition of 

18 oxides of nitrogen and sulfur. 

19 1. Adequacy of the Current Standards for Direct Effects

20 The current secondary oxides of nitrogen and sulfur standards are intended to protect 

21 against adverse effects to public welfare. For oxides of nitrogen, the current secondary standard 
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was set identical to the primary standard31 , e.g. an annual standard set for NO2 to protect against 

2 adverse effects on vegetation from direct exposure to ambient oxides of nitrogen. For oxides of 

3 sulfur, the current secondary standard is a 3-hour standard intended to provide protection for 

4 plants from the direct foliar damage associated with atmospheric concentrations of SO2. It is 

5 appropriate to consider whether the current standards are adequate to protect against the direct 

6 effects on vegetation resulting from ambient NO2 and SO2 which were the basis for the current 

7 secondary standards. The ISA concluded that there was sufficient evidence to infer a causal 

8 relationship between exposure to SO2, NO, NO2 and PAN and injury to vegetation. Additional 

9 research on acute foliar injury has been limited and there is no evidence to suggest foliar injury

10 below the levels of the current secondary standards for oxides of nitrogen and sulfur. There is 

11 sufficient evidence to suggest that the levels of the current standards are likely adequate to 

12 protect against direct phytotoxic effects. 

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

14 This section addresses two concepts necessary to evaluate the current standards in the 

15 context of deposition related effects. First, appropriateness of the current standards is considered 

16 with regard to indicator, form, level and averaging time. This discussion centers around the 

17 ability of the current standards to evaluate and provide protection against deposition related 

18 effects that vary spatially and temporally. It includes particular emphasis on the indicators and 

19 forms of the current standards and the degree to which they are ecologically relevant with regard 

20 to deposition related effects. Second, this section evaluates the current standardsin terms of 

21 adequacy of protection. 

22 a. Appropriateness

 

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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1 The ISA has established that the major effects of concern for this review of the oxides of 

2 nitrogen and sulfur standards are associated with deposition of N and S caused by atmospheric 

3 concentrations of oxides of nitrogen and sulfur. The current standards are not directed toward 

4 depositional effects, and none of the elements of the current NAAQS – indicator, form, 

5 averaging time, and level – are suited for addressing the effects of N and S deposition. 

6 Four issues arise that call into question the ecological relevance of the structure of the 

7 current secondary standards for oxides of nitrogen and sulfur. 

8 (1) The current SO2 secondary standard (0.5 ppm SO2 over a 3-hour average) does not utilize an 

9 exposure period that is relevant for ecosystem impacts. The majority of deposition related 

10 impacts are associated with depositional loads that occur over periods of months to years. 

11 This differs significantly from exposures associated with hourly concentrations of SO2 as 

12 measured by the current secondary standard. By addressing short-term concentrations, the 

13 current SO2 secondary standard, while protective against direct foliar effects from gaseous 

14 oxides of sulfur, does not take into account the findings of effects in the ISA, which notes the 

15 relationship between annual deposition of S and acidification effects which are likely to be 

16 more severe and widespread than phytotoxic effects under current ambient conditions, and 

17 include effects from long term deposition as well as short term. Acidification is a process 

18 which occurs over time, as the ability of an aquatic system to counteract acidic inputs is 

19 reduced as natural buffers are used more rapidly than they can be replaced through geologic 

20 weathering. The relevant period of exposure for ecosystems is therefore not the exposures 

21 captured in the short averaging time of the current SO2 secondary standard. The current 

22 secondary standard for oxides of nitrogen is an annual standard (0.053 ppm averaged over 1 

23 year) and as such is more ecologically relevant. Current standards do not utilize appropriate 
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1 atmospheric indicators. NO2 and SO2 are used as the component of oxides of nitrogen and 

2 sulfur that are measured, but they do not provide a complete link to the direct effects on 

3 ecosystems from deposition of oxides of nitrogen and sulfur as they do not capture all 

4 relevant chemical species of oxidized nitrogen and oxidized sulfur that contribute to 

5 deposition. The ISA provides evidence that deposition related effects are linked with total 

6 nitrogen and total sulfur deposition, and thus all forms of oxidized nitrogen and oxidized 

7 sulfur that are deposited will contribute to effects on ecosystems. Thus, by using 

8 atmospheric NO2 and SO2 concentrations as indicators, the current standards address only a 

9 fraction of total atmospheric oxides of nitrogen and sulfur, and do not take into account the 

10 effects from deposition of total atmospheric oxides of nitrogen and sulfur. This suggests that 

11 more comprehensive atmospheric indicators should be considered in designing ecologically 

12 relevant standards. 

13 (2) Current standards reflect separate assessments of the two individual pollutants, NO2 and SO2, 

14 rather than assessing the joint impacts of deposition to ecosystems. Recognizing the role that 

15 each pollutant plays in jointly affecting ecosystem indicators, functions, and services is vital 

16 to developing a meaningful standard. The clearest example of this interaction is in 

17 assessment of the impacts of acidifying deposition on aquatic ecosystems. Acidification in 

18 an aquatic ecosystem depends on the total acidifying potential of the deposition of both N 

19 and S from both atmospheric deposition of oxides of nitrogen and sulfur as well as the inputs 

20 from other sources of N and S such as reduced nitrogen and non-atmospheric sources. It is 

21 the joint impact of the two pollutants that determines the ultimate effect on organisms within 

22 the ecosystem, and critical ecosystem functions such as habitat provision and biodiversity. 

23 Standards that are set independently are less able to account for the contribution of the other 
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1 pollutant. This suggests that interactions between oxides of nitrogen and oxides of sulfur 

2 should be a critical element of the conceptual framework for ecologically relevant standards. 

3 There are also important interactions between oxides of nitrogen and sulfur and reduced 

4 forms of nitrogen, which also contribute to acidification and nutrient enrichment. It is 

5 important that the structure of the standards address the role of reduced nitrogen in 

6 determining the ecological effects resulting from deposition of atmospheric oxides of 

7 nitrogen and sulfur. Consideration will also have to be given to total loadings as ecosystems 

8 respond to all sources of N and S.

9 (3) Current standards do not take into account variability in ecosystem sensitivity. Ecosystems 

10 are not uniformly distributed either spatially or temporally in their sensitivity to oxides of 

11 nitrogen and sulfur. Therefore, failure to account for the major determinants of variability, 

12 including geological and soil characteristics related to the sensitivity to acidification or 

13 nutrient enrichment as well as atmospheric and landscape characteristics that govern rates of 

14 deposition, may lead to standards that do not provide requisite levels of protection across 

15 ecosystems. The current structures of the standards do not address the complexities in the 

16 responses of ecosystems to deposition of oxides of nitrogen and sulfur. Ecosystems contain 

17 complex groupings of organisms that respond in various ways to the alterations of soil and 

18 water that result from deposition of nitrogen and sulfur compounds. Different ecosystems 

19 therefore respond in different ways depending on a multitude of factors that control how 

20 deposition is integrated into the system. For example, the same levels of deposition falling 

21 on limestone dominated soils have a very different effect from those falling on shallow 

22 glaciated soils underlain with granite. One system may over time display no obvious 

23 detriment while the other may experience a catastrophic loss in fish communities. This 
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1 degree of sensitivity is a function of many atmospheric factors which control rates of 

2 deposition as well as ecological factors which control how an ecosystem responds to that 

3 deposition. The current standards do not take into account spatial and seasonal variations not 

4 only in depositional loadings but also in sensitivity of ecosystems exposed to those loadings. 

5 Based on the discussion summarized above, the Policy Assessment concludes that the current 

6 secondary standards for oxides of nitrogen and oxides of sulfur are not ecologically relevant 

7 in terms of averaging time, form, level or indicator.

8 b. Adequacy of Protection

9 As described in the Policy Assessment, 2005 ambient conditions indicate that the current 

10 SO2 and NO2 secondary standards are not exceeded (US EPA, 2011, Figures 6-1 and 6-2) in 

11 locations where ecological effects have been observed. In many locations, SO2 and NO2 

12 concentrations are substantially below the levels of the secondary standards. This pattern 

13 suggests that levels of deposition and any effects on ecosystems due to deposition of oxides of 

14 nitrogen and sulfur under recent conditions are occurring even though areas meet or are below 

15 current standards. In addition, these levels based on conclusions in the REA will not decline in 

16 the future to levels below which it is reasonable to anticipate effects.

17 In determining the adequacy of the current secondary standards for oxides of nitrogen 

18 and sulfur the Policy Assessment considered the extent to which ambient deposition contributes 

19 to loadings in ecosystems. Since the last review of the secondary standard for oxides of nitrogen, 

20 a great deal of information on the contribution of atmospheric deposition associated with 

21 ambient oxides of nitrogen has become available. The REA presents a thorough assessment of 

22 the contribution of oxidized nitrogen to nitrogen deposition throughout the U.S., and the relative 

23 contributions of ambient oxidized and reduced forms of nitrogen. The REA concludes that based 
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1 on that analysis, ambient oxides of nitrogen are a significant component of atmospheric nitrogen 

2 deposition, even in areas with relatively high rates of deposition of reduced nitrogen. In 

3 addition, atmospheric deposition of oxidized nitrogen contributes significantly to total nitrogen 

4 loadings in nitrogen sensitive ecosystems. 

5 The ISA summarizes the available studies of relative nitrogen contribution and finds that 

6 in much of the U.S., oxides of nitrogen contribute from 50 to 75 percent of total atmospheric 

7 deposition relative to total reactive nitrogen that includes oxidized and reduced nitrogen species 

8 (US EPA, 2008, section 2.8.4). Although the proportion of total nitrogen loadings associated 

9 with atmospheric deposition of nitrogen varies across locations (N deposition in the eastern U.S. 

10 includes locations with greater than 9 kg N/ha-yr, and in the central U.S. high deposition 

11 locations with values on the order of 6 to 7 kg N/ha-yr), the ISA indicates that atmospheric N 

12 deposition is the main source of new anthropogenic N to most headwater streams, high elevation 

13 lakes, and low-order streams. Atmospheric N deposition contributes to the total N load in 

14 terrestrial, wetland, freshwater, and estuarine ecosystems that receive N through multiple 

15 pathways. In several large estuarine systems, including the Chesapeake Bay, atmospheric 

16 deposition accounts for between 10 and 40 percent of total nitrogen loadings (US EPA, 2008). 

17 Atmospheric concentrations of oxides of sulfur account for nearly all S deposition in the 

18 US. For the period 2004–2006, mean S deposition in the U.S. was greatest east of the 

19 Mississippi River with the highest deposition amount, 21.3 kg S/ha-yr, in the Ohio River Valley 

20 where most recording stations reported 3 year averages >10 kg S/ha-yr. Numerous other stations 

21 in the East reported S deposition >5 kg S/ha-yr. Total S deposition in the U.S. west of the 100th 

22 meridian was relatively low, with all recording stations reporting <2 kg S/ha-yr and many 

reporting <1 kg S/ha-yr. S was primarily deposited in the form of wet SO4

2−

23 followed in 
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1 decreasing order by a smaller proportion of dry SO2 and a much smaller proportion of deposition 

as dry SO4

2−

2 . 

3 As discussed throughout the REA (US EPA, 2009 and section IIB above), there are 

4 several key areas of risk that are associated with ambient concentrations of oxides of nitrogen 

5 and sulfur. As noted earlier, in previous reviews of the secondary standards for oxides of 

6 nitrogen and sulfur, the standards were designed to protect against direct exposure of plants to 

7 ambient concentrations of the pollutants. A significant shift in understanding of the effects of 

8 oxides of nitrogen and sulfur has occurred since the last reviews, reflecting the large amount of 

9 research that has been conducted on the effects of deposition of nitrogen and sulfur to 

10 ecosystems. The most significant current risks of adverse effects to public welfare are those 

11 related to deposition of oxides of nitrogen and sulfur to both terrestrial and aquatic ecosystems. 

12 These risks fall into two categories, acidification and nutrient enrichment, which were 

13 emphasized in the REA as most relevant to evaluating the adequacy of the existing standards in 

14 protecting public welfare from adverse ecological effects.

15 i. Aquatic acidification

16 The focus of the REA case studies was on determining whether deposition of sulfur and 

17 oxidized nitrogen in locations where ambient oxides of nitrogen and sulfur were at or below the 

18 current standards was resulting in acidification and related effects, including episodic 

19 acidification and mercury methylation. Based on the case studies conducted for lakes in the 

20 Adirondacks and streams in Shenandoah National Park (case studies are discussed more fully in 

21 section IIB and US EPA, 2009), there is significant risk to acid sensitive aquatic ecosystems at 

22 atmospheric concentrations of oxides of nitrogen and sulfur at or below the current standards. 

23 The REA also supports strongly a relationship between atmospheric deposition of oxides of 
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1 nitrogen and sulfur and loss of ANC in sensitive ecosystems and indicates that ANC is an 

2 excellent indicator of aquatic acidification. The REA also concludes that at levels of deposition 

3 associated with oxides of nitrogen and sulfur concentrations at or below the current standards, 

4 ANC levels are expected to be below benchmark values that are associated with significant 

5 losses in fish species richness.

6 Significant portions of the U.S. are acid sensitive, and current deposition levels exceed 

7 those that would allow recovery of the most acid sensitive lakes in the Adirondacks (US EPA, 

8 2008, Executive Summary). In addition, because of past loadings, areas of the Shenandoah are 

9 sensitive to current deposition levels (US EPA, 2008, Executive Summary). Parts of the West 

10 are naturally less sensitive to acidification and subjected to lower deposition (particularly SOx) 

11 levels relative to the eastern United States, and as such, less focus in the ISA is placed on the 

12 adequacy of the existing standards in these areas, with the exception of the mountainous areas of 

13 the West, which experience episodic acidification due to deposition. 

14 In describing the effects of acidification in the two case study areas the REA uses the 

15 approach of describing benchmarks in terms of ANC values. Many locations in sensitive areas of 

16 the U.S. have ANC levels below benchmark levels for ANC classified as severe, elevated, or 

17 moderate concern (US EPA, 2011, Figure 2-1). The average current ANC levels across 44 lakes 

18 in the Adirondack case study area is 62.1 μeq/L (moderate concern). However, 44 percent of 

19 lakes had deposition levels exceeding the critical load for an ANC of 50 μeq/L, and 28 percent of 

20 lakes had deposition levels exceeding the (higher) critical load for an ANC of 20 μeq/L (US 

21 EPA, 2009, section 4.2.4.2). This information indicates that almost half of the 44 lakes in the 

22 Adirondacks case study area are at an elevated concern level, and almost a third are at a severe 

23 concern level. These levels are associated with greatly diminished fish species diversity, and 
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1 losses in the health and reproductive capacity of remaining populations. Based on assessments of 

2 the relationship between number of fish species and ANC level in both the Adirondacks and 

3 Shenandoah areas, the number of fish species is decreased by over half at an ANC level of 20 

4 μeq/L relative to an ANC level at 100 μeq/L (US EPA, 2009, Figure 4.2-1). When extrapolated 

5 to the full population of lakes in the Adirondacks area using weights based on the EMAP 

6 probability survey (US EPA, 2009, section 4.2.6.1), 36 percent of lakes exceeded the critical load 

7 for an ANC of 50 μeq/L and 13 percent of lakes exceeded the critical load for an ANC of 20 

8 μeq/L.

9 Many streams in the Shenandoah case study area also have levels of deposition that are 

10 associated with ANC levels classified as severe, elevated, or moderate concern. The average

11 ANC under recent conditions for the 60 streams evaluated in the Shenandoah case study area is 

12 57.9 μeq/L, indicating moderate concern. However, 85 percent of these streams had recent 

13 deposition exceeding the critical load for an ANC of 50 μeq/L, and 72 percent exceeded the 

14 critical load for an ANC of 20 μeq/L. As with the Adirondacks area, this information suggests 

15 that significant numbers of sensitive streams in the Shenandoah area are at risk of adverse 

16 impacts on fish populations under recent conditions. Many other streams in the Shenandoah area 

17 are also likely to experience conditions of elevated to severe concern based on the prevalence in 

18 the area of bedrock geology associated with increased sensitivity to acidification suggesting that 

19 effects due to stream acidification could be widespread in the Shenandoah area (US EPA, 2009, 

20 section 4.2.6.2).

21 In addition to these chronic acidification effects, the ISA notes that “consideration of 

22 episodic acidification greatly increases the extent and degree of estimated effects for acidifying 

23 deposition on surface waters.” (US EPA, 2008, section 3.2.1.6) Some studies show that the 
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1 number of lakes that could be classified as acid-impacted based on episodic acidification is 2 to 3 

2 times the number of lakes classified as acid-impacted based on chronic ANC. These episodic 

3 acidification events can have long term effects on fish populations (US EPA, 2008, section 

4 3.2.1.6). Under recent conditions, episodic acidification has been observed in locations in the 

5 eastern U.S. and in the mountainous western U.S. (US EPA, 2008, section 3.2.1.6). 

6 The ISA, REA and Policy Assessment all conclude that the current standards are not 

7 adequate to protect against the adverse impacts of aquatic acidification on sensitive ecosystems. 

8 In the ISA it is noted that significant portions of the U.S. are acid sensitive, and that current 

9 deposition levels exceed those that would allow recovery of the most acid sensitive lakes in the 

10 Adirondacks (US EPA, 2008, Executive Summary). In addition, because of past loadings, areas 

11 of the Shenandoah are sensitive to current deposition levels (US EPA, 2008, Executive 

12 Summary). A recent survey, as reported in the ISA, found sensitive streams in many locations in 

13 the U.S., including the Appalachian Mountains, the Coastal Plain, and the Mountainous West 

14 (US EPA, 2008, section 4.2.2.3). In these sensitive areas, between 1 and 6 percent of stream 

15 kilometers are chronically acidified. The REA further concludes that both the Adirondack and 

16 Shenandoah case study areas are currently receiving deposition from ambient oxides of nitrogen 

17 and sulfur in excess of their ability to neutralize such inputs. In addition based on the current 

18 emission scenarios, forecast modeling out to the year 2020 as well as 2050 indicates a large 

19 number of streams in these areas will still be adversely impacted (section IIB). Based on these 

20 considerations, the Policy Assessment concludes that the current secondary NAAQS for oxides 

21 of nitrogen and sulfur do not provide adequate protection of sensitive ecosystems with regard to 

22 aquatic acidification. 

23 ii. Terrestrial acidification
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1 Based on the terrestrial acidification case studies (Kane Experimental Forest in 

2 Pennsylvania and Hubbard Brook Experimental Forest described in section IIB) on sugar maple 

3 and red spruce habitat, the REA concludes that there is significant risk to sensitive terrestrial 

4 ecosystems from acidification at atmospheric concentrations of NOx and SOX at or below the 

5 current standards. The ecological indicator selected for terrestrial acidification is the base 

6 cation to aluminum ratio (BC:Al), which has been linked to tree health and growth. The results 

7 of the REA strongly support a relationship between atmospheric deposition of oxides of nitrogen 

8 and sulfur and BC:Al, and that BC:Al is a good indicator of terrestrial acidification. At levels of 

9 deposition associated with oxides of nitrogen and sulfur concentrations at or below the current 

10 standards, BC:Al levels are expected to be below benchmark values that are associated with 

11 significant effects on tree health and growth. Such degradation of terrestrial ecosystems could 

12 affect ecosystem services such as habitat provisioning, endangered species, goods production 

13 (timber, syrup, etc.) among others. 

14 Many locations in sensitive areas of the U.S. have BC:Al levels below benchmark levels 

15 classified as providing low to intermediate levels of protection to tree health. At a BC:Al ratio of 

16 1.2 (intermediate level of protection), red spruce growth can be reduced by 20 percent. At a 

17 BC:Al ratio of 0.6 (low level of protection), sugar maple growth can be decreased by 20 percent. 

18 The REA did not evaluate broad sensitive regions. However, in the sugar maple case study area 

19 (Kane Experimental Forest), recent deposition levels are associated with a BC:Al ratio below 

20 1.2, indicating between intermediate and low level of protection, which would indicate the 

21 potential for a greater than 20 percent reduction in growth. In the red spruce case study area 

22 (Hubbard Brook Experimental Forest), recent deposition levels are associated with a BC:Al ratio 
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1 slightly above 1.2, indicating slightly better than an intermediate level of protection (US EPA, 

2 2009, section 4.3.5.1). 

3 Over the full range of sugar maple, 12 percent of evaluated forest plots exceeded the 

4 critical loads for a BC:Al ratio of 1.2, and 3 percent exceeded the critical load for a BC:Al ratio 

5 of 0.6. However, there was large variability across states. In New Jersey, 67 percent of plots 

6 exceeded the critical load for a BC:Al ratio of 1.2, while in several states on the outskirts of the 

7 range for sugar maple (e.g. Arkansas, Illinois) no plots exceeded the critical load for a BC:Al 

8 ratio of 1.2. For red spruce, overall 5 percent of plots exceeded the critical load for a BC:Al ratio 

9 of 1.2, and 3 percent exceeded the critical load for a BC:Al ratio of 0.6. In the major red spruce 

10 producing states (Maine, New Hampshire, and Vermont), critical loads for a BC:Al ratio of 1.2 

11 were exceeded in 0.5, 38, and 6 percent of plots.

12 The ISA, REA and Policy Assessment all conclude that the current standards are not 

13 adequate to protect against the adverse impacts of terrestrial acidification on sensitive 

14 ecosystems. The ISA reported one study that estimated 15 percent of U.S. forest ecosystems 

15 exceeded the critical loads for acidity for N and S deposition by >250 eq/ha/year under current 

16 conditions (US EPA, 2008, section 4.2.1.3). The main drawback, as stated in the REA and 

17 Policy Assessment, with the understanding of terrestrial acidification lies in the availability of 

18 data by which we can predict critical loads and that the data are based on laboratory responses 

19 rather than field measurements. Other stressors that are present in the field but that are not 

20 present in the laboratory may confound this relationship. The REA does however, conclude that 

21 the case study results suggest that nitrogen and sulfur acidifying deposition in the sugar maple 

22 and red spruce forest areas studied were very close to, if not greater than, the critical loads for 
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1 those areas, and both ecosystems are likely to be sensitive to any future changes in the levels of 

2 deposition.

3 iii. Terrestrial nutrient enrichment

4 Nutrient enrichment effects are due to nitrogen loadings from both atmospheric and non5 atmospheric sources. Evaluation of nutrient enrichment effects requires an understanding that 

6 nutrient inputs are essential to ecosystem health and that specific long term levels of nutrients in 

7 a system affect the types of species that occur over long periods of time. Short term additions of 

8 nutrients can affect species competition, and even small additions of nitrogen in areas that are 

9 traditionally nutrient poor can have significant impacts on productivity as well as species 

10 composition. Most ecosystems in the United States are nitrogen-limited, so regional decreases 

11 in emissions and deposition of airborne nitrogen compounds could lead to some decrease in 

12 growth of the vegetation that surrounds the targeted aquatic system but as discussed below 

13 evidence for this is mixed. Whether these changes in plant growth are seen as beneficial or 

14 adverse will depend on the nature of the ecosystem being assessed. 

15 Information on the effects of changes in nitrogen deposition on forestlands and other 

16 terrestrial ecosystems is very limited. The multiplicity of factors affecting forests, including other 

17 potential stressors such as ozone, and limiting factors such as moisture and other nutrients, 

18 confound assessments of marginal changes in any one stressor or nutrient in forest ecosystems. 

19 The ISA notes that only a fraction of the deposited nitrogen is taken up by the forests, most of 

20 the nitrogen is retained in the soils (US EPA, 2008, section 3.3.2.1). In addition, the ISA 

21 indicates that forest management practices can significantly affect the nitrogen cycling within a 

22 forest ecosystem, and as such, the response of managed forests to NOx deposition will be 

23 variable depending on the forest management practices employed in a given forest ecosystem 
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1 (US EPA, 2008, Annex C C.6.3). Increases in the availability of nitrogen in N-limited forests 

2 via atmospheric deposition could increase forest production over large non-managed areas, but 

3 the evidence is mixed, with some studies showing increased production and other showing little 

4 effect on wood production (US EPA, 2008, section 3.3.9). Because leaching of nitrate can 

5 promote cation losses, which in some cases create nutrient imbalances, slower growth and 

6 lessened disease and freezing tolerances for forest trees, the net effect of increased N on forests 

7 in the U.S. is uncertain (US EPA, 2008, section 3.3.9).

8 The scientific literature has many examples of the deleterious effects caused by excessive 

9 nitrogen loadings to terrestrial systems. Several studies have set benchmark values for levels of 

10 N deposition at which scientifically adverse effects are known to occur. Large areas of the 

11 country appear to be experiencing deposition above these benchmarks. The ISA indicates that 

12 studies have found that at 3.1 kg N/ha-yr, the community of lichens begins to change from 

13 acidophytic to tolerant species; at 5.2 kg N/ha-yr, the typical dominance by acidophytic species 

14 no longer occurs; and at 10.2 kg N/ha-yr, acidophytic lichens are totally lost from the 

15 community. Additional studies in the Colorado Front Range of the Rocky Mountain National 

16 Park support these findings. These three values (3.1, 5.2, and 10.2 kg/ha-yr) are one set of 

17 ecologically meaningful benchmarks for the mixed conifer forest (MCF) of the pacific coast 

18 regions. Nearly all of the known sensitive communities receive total nitrogen deposition levels 

19 above the 3.1 N kg/ha-yr ecological benchmark according to the12 km, 2002 CMAQ/NADP 

20 data, with the exception of the easternmost Sierra Nevadas. MCFs in the southern portion of the 

21 Sierra Nevada forests and nearly all MCF communities in the San Bernardino forests receive 

22 total nitrogen deposition levels above the 5.2 N kg/ha-yr ecological benchmark. 
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1 Coastal Sage Scrub communities (CSS) are also known to be sensitive to community

2 shifts caused by excess nitrogen loadings. Studies have investigated the amount of nitrogen 

3 utilized by healthy and degraded CSS systems. In healthy stands, the authors estimated that 3.3 

4 kg N/ha-yr was used for CSS plant growth. It is assumed that 3.3 kg N/ha-yr is near the point 

5 where nitrogen is no longer limiting in the CSS community. Therefore, this amount can be 

6 considered an ecological benchmark for the CSS community. The majority of the known CSS 

7 range is currently receiving deposition in excess of this benchmark. Thus, the REA concludes 

8 that recent conditions where oxides of nitrogen ambient concentrations are at or below the 

9 current oxides of nitrogen secondary standards are not adequate to protect against anticipated 

10 adverse impacts from N nutrient enrichment in sensitive ecosystems. 

11 iv. Aquatic nutrient enrichment

12 The REA aquatic nutrient enrichment case studies focused on coastal estuaries and 

13 revealed that while current ambient loadings of atmospheric oxides of nitrogen are contributing 

14 to the overall depositional loading of coastal estuaries, other non-atmospheric sources are 

15 contributing in far greater amounts in total, although atmospheric contributions are as large as 

16 some other individual source types. The ability of current data and models to characterize the 

17 incremental adverse impacts of nitrogen deposition is limited, both by the available ecological 

18 indicators, and by the inability to attribute specific effects to atmospheric sources of nitrogen. 

19 The REA case studies used as the ecological indicator for aquatic nutrient enrichment an index 

20 of eutrophication known as the Assessment of Estuarine Trophic Status Eutrophication Index 

21 (ASSETS EI). This index is a six level index characterizing overall eutrophication risk in a 

22 waterbody. This indictor is not sensitive to changes in nitrogen deposition within a single level 
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1 of the index. In addition, this type of indicator does not reflect the impact of nitrogen deposition 

2 in conjunction with other sources of nitrogen. 

3 Based on the above considerations, the REA concludes that the ASSETS EI is not an 

4 appropriate ecological indicator for estuarine aquatic eutrophication and that additional analysis 

5 is required to develop an appropriate indicator for determining the appropriate levels of 

6 protection from N nutrient enrichment effects in estuaries related to deposition of oxides of 

7 nitrogen. As a result, EPA is unable to make a determination as to the adequacy of the existing 

8 secondary oxides of nitrogen standard in protecting public welfare from N nutrient enrichment 

9 effects in estuarine aquatic ecosystems.

10 Additionally, nitrogen deposition can alter species composition and cause eutrophication 

11 in freshwater systems. In the Rocky Mountains, for example, deposition loads of 1.5 to 2 kg/ha12 yr which are well within current ambient levels are known to cause changes in species 

13 composition in diatom communities indicating impaired water quality (US EPA, 2008, section 

14 3.3.5.3). This suggests that the existing secondary standard for oxides of nitrogen may not 

15 protect such ecosystems and their resulting services from impairment.

16 v. Other effects

17 An important consideration in looking at the effects of deposition of oxides of sulfur in 

18 aquatic ecosystems is the potential for production of methylmercury, a neurotoxic contaminant. 

The production of meaningful amounts of methylmercury (MeHg) requires the presence of SO4

2-

19

and mercury, and where mercury is present, increased availability of SO4

2-

20 results in increased 

21 production of MeHg. There is increasing evidence on the relationship between sulfur deposition 

22 and increased methylation of mercury in aquatic environments; this effect occurs only where 

23 other factors are present at levels within a range to allow methylation. The production of 
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1 methylmercury requires the presence of sulfate and mercury, but the amount of methylmercury

2 produced varies with oxygen content, temperature, pH, and supply of labile organic carbon (US 

3 EPA, 2008, section 3.4). In watersheds where changes in sulfate deposition did not produce an 

4 effect, one or several of those interacting factors were not in the range required for meaningful 

5 methylation to occur (US EPA, 2008, section 3.4). Watersheds with conditions known to be 

6 conducive to mercury methylation can be found in the northeastern United States and 

7 southeastern Canada (US EPA, 2009, section 6). 

8 With respect to sulfur deposition and mercury methylation, the final ISA determined: The 

9 evidence is sufficient to infer a causal relationship between sulfur deposition and increased 

10 mercury methylation in wetlands and aquatic environments. However, EPA did not conduct a 

11 quantitative assessment of the risks associated with increased mercury methylation under current 

12 conditions. As such, EPA is unable to make a determination as to the adequacy of the existing 

13 SO2 secondary standards in protecting against welfare effects associated with increased mercury

14 methylation.

15 vi. Summary of Adequacy Considerations

16 In summary, the Policy Assessment concludes that currently available scientific evidence 

17 and assessments clearly call into question the adequacy of the current standards with regard to 

18 deposition-related effects on sensitive aquatic and terrestrial ecosystems, including acidification 

19 and nutrient enrichment. Further, the Policy Assessment recognizes that the elements of the 

20 current standards -- indicator, averaging time, level and form – are not ecologically relevant, and 

21 are thus not appropriate for standards designed to provide such protection. Thus, the Policy

22 Assessment concludes that consideration should be given to establishing a new ecologically

23 relevant multi-pollutant, multimedia standard to provide appropriate protection from deposition-
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1 related ecological effects of oxides of nitrogen and sulfur on sensitive ecosystems, with a focus 

2 on protecting against adverse effects associated with acidifying deposition in sensitive aquatic 

3 ecosystems.

4 3. CASAC Views

5 In a letter to the Administrator (Russell and Samet 2011a), the CASAC Oxides of 

6 Nitrogen and Oxides of Sulfur Panel, with full endorsement of the chartered CASAC, 

7 unanimously concluded that:

8 EPA staff has demonstrated through the Integrated Science Assessment (ISA), 

9 Risk and Exposure Characterization (REA) and the draft PA that ambient NOx 

10 and SOx can have, and are having, adverse environmental impacts. The Panel 

11 views that the current NOx and SOx secondary standards should be retained to 

12 protect against direct adverse impacts to vegetation from exposure to gas phase 

13 exposures of these two families of air pollutants. Further, the ISA, REA and draft 

14 PA demonstrate that adverse impacts to aquatic ecosystems are also occurring due 

15 to deposition of NOx and SOx. Those impacts include acidification and 

16 undesirable levels of nutrient enrichment in some aquatic ecosystems. The levels 

17 of the current NOx and SOx secondary NAAQS are not sufficient, nor the forms 

18 of those standards appropriate, to protect against adverse depositional effects; thus 

19 a revised NAAQS is warranted.

20

21 In addition, with regard to the joint consideration of both oxides of nitrogen and oxides 

22 of sulfur as well as the consideration of deposition related effects, CASAC concluded that the 

23 Policy Assessment had developed a credible methodology for considering such effects. The 

24 Panel stated that “the Policy Assessment develops a framework for a multi-pollutant, multimedia 

25 standard that is ecologically relevant and reflects the combined impacts of these two pollutants as 

26 they deposit to sensitive aquatic ecosystems.”

27 4. Administrator’s Proposed Conclusions Concerning Adequacy of Current Standard

28 Based on the above considerations and taking into account CASAC advice, the 

29 Administrator recognizes that the purpose of the secondary standard is to protect against 

30 “adverse” effects resulting from exposure to oxides of nitrogen and sulfur, discussed above in 
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1 section IIA. The Administrator also recognizes the need for conclusions as to the adequacy of the 

2 current standards for both direct and deposition related effects as well as conclusions as to the 

3 appropriateness and ecological relevance of the current standards. 

4 In considering what constitutes an ecological effect that is also adverse to the public 

5 welfare, the Administrator took into account the ISA conclusions regarding the nature and 

6 strength of the effects evidence, the risk and exposure assessment results, the degree to which the 

7 associated uncertainties should be considered in interpreting the results, the conclusions 

8 presented in the Policy Assessment, and the views of CASAC and members of the public. On 

9 these bases, the Administrator concludes that the current secondary standards are adequate to 

10 protect against direct phytotoxic effects on vegetation. Thus, the Administrator proposes to 

retain the current secondary standard for oxides of nitrogen at 53 ppb,411 annual average 

12 concentration, measured in the ambient air as NO2, and the current secondary standard for oxides 

13 of sulfur at 0.5 ppm, 3-hour average concentration, measured in the ambient air as SO2. 

14 With regard to deposition-related effects, the Administrator has first to consider the 

15 appropriateness of the structure of the current standards to address ecological effects of concern. 

16 Based on the evidence as well as considering the advice given by CASAC on this matter, the 

17 Administrator concludes that the elements of the current standards are not ecologically relevant 

18 and thus are not appropriate to provide protection of ecosystems. On the subject of adequacy of 

19 protection with regard to deposition-related effects, the Administrator considered the full nature 

20 of ecological effects related to the deposition of ambient oxides of nitrogen and sulfur into 

21 sensitive ecosystems across the country. Her conclusions are based on the evidence presented in 

22 the ISA with regard to acidification and nutrient enrichment effects, the findings of the REA with 

 

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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1 regard to scope and severity of the current and likely future effects of deposition, the synthesis of 

2 both the scientific evidence and risk and exposure results in the Policy Assessment as to the 

3 adequacy of the current standards, and the advice of both CASAC and the public. After such 

4 consideration, the Administrator concludes that current levels of oxides of nitrogen and sulfur are 

5 sufficient to cause acidification of both aquatic and terrestrial ecosystems, nutrient enrichment of 

6 terrestrial ecosystems and contribute to nutrient enrichment effects in estuaries that could be 

7 considered adverse, and the current secondary standards do not provide adequate protection from 

8 such effects. 

9 Having reached these conclusions, the Administrator determines that it is appropriate to 

10 consider alternative standards that are ecologically relevant. These considerations support the 

11 conclusion that the current secondary standards is neither appropriate nor adequate to protect 

12 against deposition related effects. The Administrator’s consideration of such alternative 

13 standards is discussed below in Section III.

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

15 Secondary Standards for Aquatic Acidification

16 Having reached the conclusion that the current NO2 and SO2 secondary standards are not 

17 adequate to provide appropriate protection against deposition-related effects associated with 

18 oxides of nitrogen and sulfur, the Administrator then considered what new multi-pollutant 

19 standard might be appropriate, at this time, to address such effects on public welfare. The 

20 Administrator recognizes that the inherently complex and variable linkages between ambient 

21 concentrations of nitrogen and sulfur oxides, the related deposited forms of nitrogen and sulfur, 

22 and the ecological responses that are associated with public welfare effects call for consideration 

23 of an ecologically relevant design of a standard that reflects these linkages. The Administrator 
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1 also recognizes that characterization of such complex and variable linkages will necessarily 

2 require consideration of information and analyses that have important limitations and 

3 uncertainties.

4 Despite its complexity, an ecologically relevant multi-pollutant standard to address 

5 deposition-related effects could still appropriately be defined in terms of the same basic elements 

6 that are used to define any NAAQS – indicator, form, averaging time, and level. The form 

7 would incorporate additional structural elements that reflect relevant multi-pollutant and 

8 multimedia attributes. These structural elements include the use of an ecological indicator, tied 

9 to the ecological effect we are focused on, and other elements that account for ecologically

10 relevant factors other than ambient air concentrations. All of these elements would be needed to 

11 enable a linkage from ambient air indicators to the ecological indicator to define an ecologically 

12 relevant standard. As a result, such a standard would necessarily be more complex than the 

13 NAAQS that have been set historically to address effects associated with ambient concentrations 

14 of a single pollutant.

15 More specifically, the Administrator considered an ecologically relevant multi-pollutant

16 standard to address effects associated with acidifying deposition related to ambient 

17 concentrations of oxides of nitrogen and sulfur in sensitive aquatic ecosystems. This focus is 

18 consistent with the information presented in the ISA, REA, and Policy Assessment which 

19 highlighted the sufficiency of the quantity and quality of the available evidence and assessments 

20 associated with aquatic acidification relative to the information and assessments available for 

21 other deposition-related effects, including terrestrial acidification and aquatic and terrestrial 

22 nutrient enrichment. Based on its review of these documents, CASAC agreed that aquatic 

23 acidification should be the focus for developing a new multi-pollutant standard in this review. In 
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1 reaching conclusions about an air quality standard designed to address deposition-related aquatic 

2 acidification effects, the Administrator also recognizes that such a standard may also provide 

3 some degree of protection against other deposition-related effects.

4 As discussed in chapter 7 of the Policy Assessment, the development of a new multi5 pollutant standard to address deposition-related aquatic acidification effects recognizes the need 

6 for consideration of a nationally applicable standard for protection against adverse effects of 

7 aquatic acidification to public welfare, while recognizing the complex and heterogeneous 

8 interactions between ambient air concentrations of nitrogen and sulfur oxides, the related 

9 deposition of nitrogen and sulfur, and associated ecological responses. The development of such 

10 a standard also needs to take into account the limitations and uncertainties in the available 

11 information and analyses upon which characterization of such interactions are based. The 

12 approach used in the Policy Assessment also recognizes that while such a standard would be 

13 national in scope and coverage, the effects to public welfare from aquatic acidification will not 

14 occur to the same extent in all locations in the U.S., given the inherent variability of the 

15 responses of aquatic systems to the effects of acidifying deposition.

16 As discussed above in section II, many locations in the U.S. are naturally protected 

17 against acid deposition due to underlying geological conditions. Likewise, some locations in the 

18 U.S., including lands managed for commercial agriculture and forestry, are not likely to be 

19 negatively impacted by current levels of nitrogen and sulfur deposition. As a result, while a new 

20 ecologically relevant secondary standard would apply everywhere, it would be structured to 

21 account for differences in the sensitivity of ecosystems across the country. This would allow for 

22 appropriate protection of sensitive aquatic ecosystems, which are relatively pristine and wild and 
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1 generally in rural areas, and the services provided by such sensitive ecosystems, without 

2 requiring more protection than is needed elsewhere. 

3 As discussed below, the multi-pollutant standard developed in the Policy Assessment 

4 would employ (1) NOy and SOx as the atmospheric ambient air indicators; (2) a form that takes 

5 into account variable factors, such as atmospheric and ecosystem conditions that modify the 

6 amounts of deposited nitrogen and sulfur; the distinction between oxidized and reduced forms of 

7 nitrogen; effects of deposited nitrogen and sulfur on aquatic ecosystems in terms of the 

8 ecological indicator ANC; and the representativeness of water bodies within a defined spatial 

9 area; (3) a multi-year averaging time, and (4) a standard level defined in terms of a single, 

10 national target ANC value that, in the context of the above form, identifies the levels of 

11 concentrations of NOy and SOx in the ambient air that would meet the standard. The form of 

12 such a standard has been defined by an index, termed an aquatic acidification index (AAI), 

13 which reflects the relationship between ambient concentrations of NOy and SOx and aquatic 

14 acidification effects that result from nitrogen and sulfur deposition related to these ambient 

15 concentrations.

16 In presenting the considerations associated with such an air quality standard to address 

17 deposition-related aquatic acidification effects, the following sections focus on each element of 

18 the standard, including indicator (section III.A), form (section III.B), averaging time (section 

19 III.C), and level (section III.D). Alternative combinations of levels and forms are discussed in 

20 section III.E. Considerations related to important uncertainties inherent in such an approach are 

21 discussed in section III.F. Advice from CASAC on such a new standard is presented in section 

22 III.G. The Administrator’s proposed decisions on such a new standard are presented in section 

23 III.H.
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1 A. Ambient Air Indicators

2 In considering alternative ambient air indicators, the Policy Assessment primarily focuses 

3 on the important attribute of association. Association in a broad sense refers to how well an 

4 ambient air indicator relates to the ecological effects of interest by virtue of both the framework 

5 that links the ambient indicator and effects and the empirical evidence that quantifies the 

6 linkages. The Policy Assessment also considers how measurable or quantifiable an indicator is 

7 to enable its use as an effective indicator of relevant ambient air concentrations.

8 As discussed above in section II.C, the Policy Assessment concludes that indicators other 

9 than NO2 and SO2 should be considered as the appropriate indicators of oxides of nitrogen and 

10 sulfur in the ambient air for protection against the acidification effects associated with deposition 

11 of the associated nitrogen and sulfur. This conclusion is based on the recognition that all forms 

12 of nitrogen and sulfur in the ambient air contribute to deposition and resulting acidification, and 

13 as such, NO2 and SO2 are incomplete indicators. In principle, the ambient indicators should 

14 represent the species that are associated with oxides of nitrogen and sulfur in the ambient air and 

15 can contribute acidifying deposition. This includes both the species of oxides of nitrogen and 

16 sulfur that are directly emitted as well as species transformed in the atmosphere from oxides of 

17 nitrogen and sulfur that retain the nitrogen and sulfur atoms from directly emitted oxides of 

18 nitrogen and sulfur. All of these compounds are associated with oxides of nitrogen and sulfur in 

19 the ambient air and can contribute to acidifying deposition. 

20 The Policy Assessment focuses in particular on the various compounds with nitrogen or 

21 sulfur atoms that are associated with oxides of nitrogen and sulfur, because the acidifying 

22 potential is specific to nitrogen and sulfur, and not other atoms (e.g., H, C, O) whether derived 

23 from the original source of oxides of nitrogen and sulfur emissions or from atmospheric 
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1 transformations. For example, the acidifying potential of each molecule of NO2, NO, HNO3 or 

2 PAN is identical, as is the potential for each molecule of SO2 or ion of particulate sulfate, p-SO4. 

3 Each atom of sulfur affords twice the acidifying potential of each atom of nitrogen.

4 1. Oxides of Sulfur

5 As discussed in the Policy Assessment (US EPA, 2011, section 7.1.1), oxides of sulfur 

6 include the gases sulfur monoxide (SO), sulfur dioxide (SO2), sulfur trioxide (SO3), disulfur 

7 monoxide (S2O), and particulate-phase sulfur compounds that result from gas-phase sulfur 

8 oxides interacting with particles. However, the sum of SO2 and SO4 does represent virtually the 

9 entire ambient air mass of sulfur that contributes to acidification. In addition to accounting for 

10 virtually all the potential for acidification from oxidized sulfur in the ambient air, there are 

11 reliable methods to monitor the concentrations of SO2 and particulate SO4. In addition, much of 

12 the data used to develop the technical basis for the standard developed in the Policy Assessment 

is based on monitoring or modeling of these species.513 The Policy Assessment concludes that the 

14 sum of SO2 and SO4, referred to as SOx, are appropriate ambient air indicators of oxides of 

15 sulfur because they represent virtually all of the acidification potential of ambient air oxides of 

16 sulfur and there are reliable methods suitable for measuring SO2 and SO4.

17 2. Oxides of Nitrogen

18 As discussed in the Policy Assessment (US EPA, 2011,section 7.1.2), total reactive 

19 oxidized nitrogen, NOy, as defined in chapter 2 of the Policy Assessment, incorporates basically 

20 all of the oxidized nitrogen species that have acidifying potential and as such, NOy should be 

21 considered as an appropriate indicator for oxides of nitrogen. NOy is an aggregate measure of

 

5 As discussed in chapter 2 of the Policy Assessment, SO2 and particulate SO4 are routinely measured in ambient air 

monitoring networks, although only CASTNET filter packs do not intentionally exclude particle size fractions. The 

CMAQ treatment of SOx is the simple addition of both species, which are treated explicitly in the model

formulation. All particle size fractions are included in the CMAQ SOx estimates.
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1 NO and NO2 and all of the reactive oxidized products of NO and NO2. That is, NOy is a group 

2 of nitrogen compounds in which all of the compounds are either an oxide of nitrogen or 

3 compounds in which the nitrogen atoms came from oxides of nitrogen. NOy is especially 

4 relevant as an ambient indicator for acidification in that it both relates to the oxides of nitrogen in 

5 the ambient air and also represents the acidification potential of all oxidized nitrogen species in 

6 the ambient air, whether an oxide of nitrogen or derived from oxides of nitrogen. 

7 There are currently available reliable methods of measuring aggregate NOy. The term 

8 “aggregate” measure means that the NOY, as measured, is not based on measuring each 

9 individual species of NOy and calculating an NOy value by summing the individual species. 

10 Rather, as described in chapter 2 of the Policy Assessment, current measurement techniques 

11 process all of the individual NOy species to produce a single aggregate measure of all of the 

12 nitrogen atoms associated with any NOy species. Consequently, the NOy measurement 

13 effectively provides the sum of all individual species, but the identity of the individual species is 

14 lost. As discussed above, the accounting for the individual nitrogen atoms is an accounting of 

15 the ambient air acidification potential of oxides of nitrogen and their transformation products and 

16 therefore the most relevant ambient indicator for aquatic acidification effects associated with 

17 oxides of nitrogen. 

18 This loss of the information on individual species motivated consideration of alternative 

19 or more narrowly defined indicators for oxides of nitrogen in the Policy Assessment. 

20 Consideration of a subset of NOy species was based on the following reasoning. First, the actual 

21 dry deposition of nitrogen is determined on an individual species basis by multiplying the species 

22 concentration times a species-specific deposition velocity and then summed to develop an 

23 estimate of total dry deposition. Consequently, the use of individual ambient species has the 
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1 potential to be more consistent with the underlying science of deposition and, therefore, has the 

2 potential to allow for a more rigorous evaluation of dry deposition with specialized field studies. 

3 In addition, there has been a suggestion of focusing only on the most quickly depositing NOy 

4 species, such as HNO3, as contributions from other NOy species such as NO2 may be negligible. 

5 These alternative indicators are discussed below.

6 The Policy Assessment considers the relative merits of using each individual NOy

7 species as part of a group of indicators. In so doing, it was first noted that dry deposition of NOy

8 is treated as the sum of the deposition of each individual species in advanced process-based air 

9 quality models like CMAQ, as described in chapter 2 of the Policy Assessment. Conceptually

10 one could extend this process-based approach by using all NOY species individually as separate 

11 indicators for oxides of nitrogen and requiring, for example, measurements of each of the 

12 species, including the dominant species of HNO3, particulate nitrate (p-NO3), true NO2, NO, and 

13 PAN. The potential attraction of using individual species would be the reliance on actual 

14 deposition velocities. This could have more physical meaning in comparison to a constructed 

15 model of aggregate deposition of NOy, which is difficult to evaluate with observations because 

16 of the assimilation of many species with disparate deposition behavior. The Policy Assessment 

17 notes that the major drawback of using individual NOy species as the indicators is the lack of 

18 reliable measurement techniques, especially for PAN and NO2 in rural locations, which renders 

19 the use of virtually any individual NOy species, except for NO and perhaps p-NO3, as 

20 functionally inadequate from a measurement perspective.

21 The Policy Assessment next considered the relative merits of using a subset of NOy

22 species as the indicators for oxides of nitrogen, as was discussed above for oxides of sulfur. To 

23 the extent that certain species provide relatively minor contributions to total NOy deposition, it 
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1 may be appropriate to consider excluding them as part of the indicator. As discussed in chapter 2 

2 of the Policy Assessment, each nitrogen species within the array of NOy species has species3 specific dry deposition velocities. For example, the deposition velocity of HNO3 is much 

4 greater than the velocity for nitrogen dioxide and, consequently, for a similar ambient air 

5 concentration, HNO3 contributes more deposition of acidifying nitrogen relative to nitrogen 

6 dioxide. In transitioning from source-oriented urban locations to rural environments, the ratio of 

7 the concentrations of HNO3 and PAN to NO2 increases.

8 Based on the reasoning that a larger fraction of the deposited NOY is accounted for by

9 total nitrate (the sum of HNO3 and p-NO3), a surrogate for the more rapidly depositing fraction 

10 of NOy, combined with the availability of reliable total nitrate measurements through the Clean 

11 Air Status and Trends Network (CASTNET), the Policy Assessment considered using total 

12 nitrate as the indicator for oxides of nitrogen (US EPA, 2011, appendix E). Nitrate would be 

13 expected to correlate well with total reactive oxidized nitrogen deposition relative to NOY (US 

14 EPA, 2011,chapter 2), given the inherent noise associated with variable contributions of low 

15 deposition velocity species (e.g., NO2) that may have relatively high ambient concentrations. 

16 However, modeling simulations suggest that NOy may be a more robust indicator, relative to 

17 HNO3, in terms of relating absolute changes in ambient air concentrations to changes in nitrogen 

18 deposition driven by changes in ambient concentrations of oxides of nitrogen (US EPA, 

19 2011,Figure 2-32).

20 Based on the above considerations, the Policy Assessment concludes that total reactive 

21 oxidized nitrogen, NOy, should be considered as the appropriate ambient indicator for oxides of 

22 nitrogen based on its direct relationship to oxides of nitrogen in the ambient air and its direct 

23 relationship to deposition associated with aquatic acidification. Because NOy represents all of 
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1 the potentially acidifying oxidized nitrogen species in the ambient air, it is appropriately 

2 associated with the deposition of potentially acidifying compounds associated with oxides of 

3 nitrogen in the ambient air. In addition, there are reliable methods available to measure NOy. 

4 Measurement of each individual species of NOy, or the measurement of only a subset of species 

5 of NOy, is less appropriate because there are not reliable measurements methods available to 

6 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 relates to acidification.

6

7

8 B. Form

9 Based on the evidence of the aquatic acidification effects caused by the deposition of 

10 NOy and SOx, the Policy Assessment (US EPA, 2011,section 7.2) presents the development of a 

11 new form that is ecologically relevant for addressing such effects. The conceptual design for the 

12 form of such a standard includes three main components: an ecological indicator, deposition 

13 metrics that relate to the ecological indicator, and a function that relates ambient air indicators to 

14 deposition metrics. Collectively, these three components link the ecological indicator to ambient 

15 air indicators, as illustrated above in Fig II-1.

16 The simplified flow diagram in Figure II-1 compresses the various atmospheric, 

17 biological, and geochemical processes associated with acidifying deposition to aquatic 

18 ecosystems into a simplified conceptual picture. The ecological indicator (left box) is related to 

19 atmospheric deposition through biogeochemical ecosystem models (middle box), which 

20 associate a target deposition load to a target ecological indicator. Once a target deposition is 

21 established, associated allowable air concentrations are determined (right box) through the 

22 relationships between concentration and deposition that are embodied in air quality models such 

 

6 The Policy Assessment also notes that NOY is a useful measurement for model evaluation purposes, which is 

especially important, recognizing the unique role that CMAQ plays in the development of this standard, as described 

below in section III.B.
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1 as CMAQ. The following discussion describes the development and rationale for each of these 

2 components, as well as the integration of these components into the full expression of the form 

3 of the standard using the concept of a national AAI that represents a target ANC level as a 

4 function of ambient air concentrations. Spatial aggregation issues associated with defining each 

5 of the terms of this index are also addressed below.

6 The AAI is designed to be an ecologically relevant form of the standard that determines 

7 the levels of NOy and SOx in the ambient air that would achieve a target ANC limit for the U.S. 

8 The intent of the AAI is to weight atmospheric concentrations of oxides of nitrogen and sulfur by 

9 their propensity to contribute to acidification through deposition, given the fundamental 

10 acidifying potential of each pollutant, and to take into account the ecological factors that govern 

11 acid sensitivity in different ecosystems. The index also accounts for the contribution of reduced 

12 nitrogen to acidification. Thus, the AAI encompasses those attributes of specific relevance to 

13 protecting ecosystems from the acidifying potential of ambient air concentrations of NOy and 

14 SOx.

15 1. Ecological Indicator

16 In considering alternative ecological indicators, the Policy Assessment again primarily

17 focuses on the attribute of association. In the case of an ecological indicator for aquatic 

18 acidification, association refers to the relationship between the indicator and adverse effects as 

19 discussed in section II. Because of the conceptual structure of the form of this standard (Figure 

20 III-1), this particular ecological indicator must also link up in a meaningful and quantifiable 

21 manner with acidifying atmospheric deposition. In effect, the ecological indicator for aquatic 

22 acidification is the bridge between biological impairment and deposition of NOy and SOx. 
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1 This section presents the rationale in the Policy Assessment for selecting acid 

2 neutralizing capacity (ANC) as the appropriate ecological indicator for consideration. 

3 Recognizing that ANC is not itself the causative or toxic agent for adverse aquatic acidification 

4 effects, the rationale for using ANC as the relevant ecological indicator is based on the 

5 following:

6 (1) ANC is directly associated with the causative agents, pH and dissolved aluminum, both 

7 through empirical evidence and mechanistic relationships;

8 (2) Empirical evidence shows very clear and strong relationships between adverse effects 

9 and ANC;

10 (3) ANC is a more reliable indicator from a modeling perspective, allowing use of a body of 

11 studies and technical analyses related to ANC and acidification to inform the 

12 development of the standard; and

13 (4) ANC literally embodies the concept of acidification as posed by the basic principles of 

14 acid base chemistry and the measurement method used to estimate ANC and, therefore, 

15 serves as a direct index to protect against acidification. 

16

17 Ecological indicators of acidification in aquatic ecosystems can be chemical or biological 

18 components of the ecosystem that are altered by the acidifying effects of nitrogen and sulfur 

19 deposition. A desirable ecological indicator for aquatic acidification is one that is measurable or 

20 estimable, linked causally to deposition of nitrogen and sulfur, and linked causally, either 

21 directly or indirectly to ecological effects known or anticipated to adversely affect public 

22 welfare.
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1 As summarized in chapter 2 of the Policy Assessment, atmospheric deposition of NOy

and SOx causes aquatic acidification through the input of strong acid anions (e.g., NO3

-

2 and 

SO4

2-

3 ) that ultimately shifts the water chemistry equilibrium toward increased hydrogen ion 

4 levels (or decreased pH). The anions are deposited either directly to the aquatic ecosystem or 

5 indirectly via transformation through soil nitrification processes and subsequent drainage from 

6 terrestrial ecosystems. In other words, when these anions are mobilized in the terrestrial soil, 

7 they can leach into adjacent water bodies. Aquatic acidification is indicated by changes in the 

8 surface water chemistry of ecosystems. In turn, the alteration of surface water chemistry has been 

9 linked to negative effects on the biotic integrity of freshwater ecosystems. There is a suite of 

10 chemical indicators that could be used to assess the effects of acidifying deposition on lake or 

11 stream acid-base chemistry. These indicators include acid neutralizing capacity (ANC); 

12 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-

13 , 

NO3

-

), cations (Ca2+, Mg+2, K+14 ), or sums of cations or anions.

15 ANC and ALK are very similar quantities and are used interchangeably in the literature 

16 and for some of the analyses presented in this document. Both ANC and ALK are defined as the 

17 amount of strong acid required to reach a specified equivalence point. For acid-base solutions, 

18 an equivalence point can be thought of as the point at which the addition of strong acids (i.e., 

19 titration) is no longer neutralized by the solution. This explains the term acid neutralizing 

20 capacity, or ANC, as ANC relates directly to the capacity of a system to neutralize acids. The 

21 differences between ANC and ALK are based on operational definitions and subject to various 

22 interpretations. ANC is preferred over ALK as the body of scientific evidence has focused on 

23 ANC and effects relationships. ALK is more widely associated with more general 
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1 characterizations of water quality such as the relative hardness of water associated with 

2 carbonates.

3 Indictors such as the concentrations of specific anions, cations, or their groupings, while 

4 relevant to acidification processes, are not robust acidification indicators as it is the relative 

5 balance of cations and anions that is more directly associated with acidification. That balance is 

6 captured by ANC and ALK. Acidity, ACY, is the coverse of ANC and indicates how much 

7 strong base it takes to reach an equivalence point. Because ACY is not used in most ecosystem 

8 assessments, the body of information relating ACY to effects is too limited to serve as a basis for 

9 an appropriate ecological indicator. Aluminum and other metals are causative toxic agents that 

10 directly impair biological functions. However, aluminum, or metals in general, have high 

11 variability in concentrations that can be linked to effects, often at extremely low levels which in 

12 some cases approach detectability limits, exhibit rapid transient responses, and are often 

13 confounded by the presence of other toxic metals. These concerns limit the use of metals as 

reliable and measurable ecological indicators. Hydrogen ion (H+14 ) concentrations, using their 

15 negative logarithmic values, or pH, are well correlated with adverse effects, as discussed above 

16 in section II.A, and determine the solubility of metals such as aluminum. However, pH is not a 

17 preferred acidification indicator due to its highly transient nature and other concerns, as 

18 discussed below. 

19 Having reasoned that ANC is a preferred indicator to ALK, ACY, individual metals or 

20 groupings of ions, the Policy Assessment considers the relative merits of ANC compared to pH, 

21 which is a well recognized indicator of acidity and a more direct causative agent with regard to 

22 adverse effects. First, the linkage between ANC and pH is considered in recognition of the 

23 causative association between pH and effects.
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1 ANC is not the direct causative toxic agent impacting aquatic species diversity. The 

2 scientific literature generally emphasizes the links between pH and adverse effects as described 

3 above in section II.A. It is important, therefore, to consider the extent to which ANC and pH are 

4 well related from a mechanistic perspective as well as through empirical evidence. ANC and 

5 pH are co-dependent on each other based on the requirement that all solutions are electrically

6 neutral, meaning that any solution must satisfy the condition that all negatively charged species 

7 must be balanced by all positively charged species. ANC is defined as the difference between 

8 strong anions and cations (US EPA, 2011, equation 7-13).

9 While the chemistry can be complex, the co-dependency between ANC and pH is 

explained by recognizing that positively charged hydrogen, H+10 , is incorporated in the charge 

11 balance relationships related to the overall solution chemistry which also defines ANC. The 

12 positive, directional co-dependency (i.e., ANC and pH increase together) is further explained in 

13 concept as ANC reflects how much strong acid (i.e., how much hydrogen ion) it takes to titrate to 

14 an equivalence point. Strong observed correlations between pH and ANC as described in the 

15 Policy Assessment support these mechanistic relationships. 

16 As discussed above in section II.A, there are well established examples of ANC 

17 correlating strongly with a variety of ecological effects which are summarized in the Policy

18 Assessment (US EPA, 2011, Table 3-1). Because pH and ANC are well correlated and linearly 

19 dependent over the pH ranges (4.5-6) where adverse ecological effects are observed, evidence of 

20 clear associations exist between ANC and adverse ecological effects as described in the Policy

21 Assessment. In large measure, this dependence between pH and ANC and the relationship of 

22 both pH and ANC to effects, speak directly to the appropriateness of ANC with respect to its use 

23 as an ecological indicator.
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1 Thus, there is a clear association between ANC and ecological effects, although there is a 

2 more direct causal relationship between pH and effects. Nonetheless, ANC is preferred as an 

3 ecological indicator based on its superior ability to provide a linkage with deposition in a 

4 meaningful and quantifiable manner, a role that is served far more effectively by ANC than by 

5 pH. While both ANC and pH are clearly associated with the effects of concern, ANC is superior 

6 in linking these effects to deposition. 

7 The Policy Assessment notes that the basis for this conclusion is that acidifying

8 atmospheric deposition of nitrogen and sulfur is a direct input of potential acidity (ACY), or, in 

9 terms of ANC, such deposition is relevant to the major anions that reduce the capacity of a water 

10 body to neutralize acidity. Consequently, there is a well defined linear relationship between 

11 potential acidifying deposition and ANC. This ANC-deposition relationship facilitates the 

12 linkage between ecosystem models that calculate an ecological indicator and the atmospheric 

13 deposition of NOy and SOx. On the other hand, there is no direct linear relationship between 

14 deposition and pH. While acid inputs from deposition lower pH, the relationship can be 

15 extremely nonlinear and there is no direct connection from a modeling or mass balance 

16 perspective between the amount of deposition entering a system and pH. The term “mass 

17 balance” underlies the basic formulation of any physical modeling construct, for atmospheric or 

18 aquatic systems, and refers to the accounting of the flow of mass into a system, the 

19 transformation to other forms, and the loss due to flow out of a system and other removal 

20 processes. ANC is a conserved property. This means that ANC in a water body can be 

21 accounted for by knowledge of how much ANC initially exists, how much flows in and is 

22 deposited, and how much flows out. In contrast, hydrogen ion concentration in the water, the 

23 basis for pH, is not a conserved property as its concentration is affected by several factors such 
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1 as temperature, atmospheric pressure, mixing conditions of a water body, and the levels of

2 several other chemical species in the system. The disadvantage of pH lacking conservative 

3 properties is that there is a very complex connection between changes in ambient air 

4 concentrations of NOy and SOx and pH. 

5 The discussion of basic water chemistry of natural systems in chapter 2 of the Policy

6 Assessment provides further details on why pH is not a conserved quantity and is subject to rapid 

7 transient response behavior that makes it difficult to use as a reliable and functional ecological 

8 indicator. The observed pH-to-ANC relationship (US EPA, 2011, figure 7-2) partially explains 

9 the concern with pH responding too abruptly. In the region where pH ranges roughly from 4.5 

10 to 6 and is of greatest relevance to effects (US EPA, 2011, figure 7-4), there clearly is more 

11 sensitivity of pH to changes in ANC in the ANC range from approximately 0 to 50 μeq/L. A 

12 focus on this part of the ANC-to-pH relationship shows that ANC associates well with pH in a 

13 fairly linear manner. However, the pH range from 4.5 to 6 also includes one of the very steepest 

14 parts of the slope relating pH as a function of ANC, where ANC ranges down below 0 μeq/L, 

15 which is subject to very rapid change in ANC, or deposition inputs. This part of the relationship 

16 coincides with reduced levels of ANC and hence with reduced ability to neutralize acids and 

17 moderate pH fluctuations. This response behavior can be extended to considering how pH 

18 would change in response to deposition, or ambient concentrations, of NOy and SOx, which can 

19 be viewed as “ANC-like” inputs. 

20 In summary, because ANC clearly links both to biological effects of aquatic acidification 

21 as well as to acidifying inputs of NOy and SOx deposition, the Policy Assessment concludes that 

22 ANC is an appropriate ecological indicator for relating adverse aquatic ecosystem effects to 

23 acidifying atmospheric deposition of SOx and NOy, and is preferred to other potential indicators. 
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1 In reaching this conclusion, the Policy Assessment notes that in its review of the first draft Policy

2 Assessment, CASAC concluded that “information on levels of ANC protective to fish and other 

3 aquatic biota has been well developed and presents probably the lowest level of uncertainty in 

4 the entire methodology” (Russell and Samet, 2010a). In its more recent review of the second 

5 draft Policy Assessment, CASAC agreed “that acid neutralizing capacity is an appropriate 

6 ecological measure for reflecting the effects of aquatic acidification” (Russell and Samet, 2010b; 

7 p. 4). 

8 2. Linking ANC to Deposition

9 There is evidence to support a quantified relationship between deposition of nitrogen and 

10 sulfur and ANC. This relationship was analyzed in the REA for two case study areas, the 

11 Adirondack and Shenandoah Mountains, based on time-series modeling and observed trends. 

12 In the REA analysis, long-term trends in surface water nitrate, sulfate and ANC were modeled 

13 using Model of Acidification of Groundwater in Catchment (MAGIC) for the two case study 

14 areas. These data were used to compare recent surface water conditions in 2006 with 

15 preindustrial conditions (i.e. preacidification 1860). The results showed a marked increase in the 

16 number of acid impacted lakes, characterized as a decrease in ANC levels, since the onset of 

17 anthropogenic nitrogen and sulfur deposition, as discussed in chapter 2 of the Policy Assessment.

18 In the REA, more recent trends in ANC, over the period from 1990 to 2006, were assessed using 

19 monitoring data collected at the two case study areas. In both case study areas, nitrate and 

20 sulfate deposition decreased over this time period. In the Adirondack Mountains, this 

21 corresponded to a decreased concentration of nitrate and sulfate in the surface waters and an 

22 increase in ANC (US EPA, 2009, section 4.2.4.2). In the Shenandoah Mountains, there was a 

23 slight decrease in nitrate and sulfate concentration in surface waters corresponding to modest 
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1 increase in ANC from 50 ueq/L in 1990 to 67 ueq/L in 2006 (US EPA, 2009, section 

2 4.2.4.3,Appendix 4, and section 3.4).

3 In the REA, the quantified relationship between deposition and ANC was investigated 

4 using ecosystem acidification models, also referred to as acid balance models or critical loads 

5 models (US EPA, 2011, section 2 and US EPA, 2009, section 4 and Appendix 4). These models 

6 quantify the relationship between deposition of nitrogen and sulfur and the resulting ANC in 

7 surface waters based on an ecosystem’s inherent generation of ANC and ability to neutralize 

8 nitrogen deposition through biological and physical processes. A critical load is defined as the 

9 amount of acidifying atmospheric deposition of nitrogen and sulfur beyond which a target ANC 

10 is not reached. Relatively high critical load values imply that an ecosystem can accommodate 

11 greater deposition levels than lower critical loads for a specific target ANC level. Ecosystem 

12 models that calculate critical loads form the basis for linking deposition to ANC. 

13 As discussed in chapter 2 of the Policy Assessment, both dynamic and steady state 

14 models calculate ANC as a function of ecosystem attributes and atmospheric nitrogen and sulfur 

15 deposition, and can be used to calculate critical loads. Steady state models are time invariant 

16 and reflect the long term consequences associated with an ecosystem reaching equilibrium under 

17 a constant level of atmospheric deposition. Dynamic models are time variant and take into 

18 account the time dependencies inherent in ecosystem hydrology, soil and biological processes. 

19 Dynamic models like MAGIC can provide the time series response of ANC to deposition 

20 whereas steady state models provide a single ANC relationship to any fixed deposition level. 

21 Dynamic models naturally are more complex than steady state models as they attempt to capture 

22 as much of the fundamental biogeochemical processes as practicable, whereas steady state 
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1 models depend on far greater parameterization and generalization of processes that is afforded, 

2 somewhat, by not having to accounting for temporal variability. 

3 The Policy Assessment notes that steady state models are capable of addressing the 

4 question of what does it take to reach and sustain a specific level of ANC. Dynamic models are 

5 also capable of addressing that question, but can also address the question of how long it takes to 

6 achieve that result. Dynamic models afford the ability for more comprehensive treatment of a 

7 variety of processes throughout the surface, soil and bedrock layers within an ecosystem. For 

8 example, steady state models treat sulfate as a mobile anion throughout the system, meaning that 

9 the sulfate that is deposited to a watershed enters the water column and is not influenced by soil 

10 adsorption or cation exchange. Dynamic models can incorporate these time variant processes. 

11 The use of a steady state model treating sulfate as totally mobile does not necessarily conflict 

12 with the possibility of sulfate acting as a less than mobile ion at certain times. The steady state 

13 assumption is premised on the long term behavior of sulfate which can undergo periods of net 

14 adsorption followed by periods of net desorption which can balance out over time. The Policy 

15 Assessment recognizes that as the richness of the available data increases, in terms of parameters 

16 and spatial resolution, the incorporation of dynamic modeling approaches in the standard setting 

17 process should become more feasible. In determining an appropriate modeling approach for the 

18 development of a NAAQS in this review, the Policy Assessment considers both the relevance of 

19 the question addressed as well as the ability to perform modeling that provides relevant 

20 information for geographic areas across the country. 

21 Dynamic models require a large amount of catchment level-specific data relative to 

22 steady state models. Because of the time invariant nature of steady state models, the data 

23 requirements that integrate across a broad spectrum of ecosystem processes is achievable and 
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1 available now at the national level. Water quality data to support steady state models currently

2 exist for developing a national data base for modeling nearly 10,000 catchments in the 

3 contiguous U.S. In contrast, the data needs to support dynamic models for national-scale 

4 analyses simply are not available at this time. Further, the information provided by steady state 

5 modeling would be sufficient to develop and analyze alternative NAAQS and the kind of 

6 protection they would afford. While it would be of interest to also obtain information about how 

7 much time it would take for a target ANC level to be achieved, the absence of such information 

8 does not preclude developing and evaluating alternative NAAQS using the AAI structure. Based 

9 on the above considerations, the Policy Assessment concludes that at this time steady state 

10 critical load modeling is an appropriate tool for linking long-term ANC levels to atmospheric 

11 deposition of nitrogen and sulfur for development of an AAI that has national applicability.

12 A steady state model is used to define the critical load, which is the amount of 

13 atmospheric deposition of nitrogen (N) and sulfur (S) beyond which a target ANC is not 

achieved and sustained.714 It is expressed as:

15

CLANClim(N + S) = ([BC]0

*

16 - [ANClim])Q + Neco (III-1)

17

18 Where:

19 CLANClim(N + S) is the critical load of deposition, with units of equivalent charge/(area-time);

[BC]0

*

20 is the natural contribution of base cations from weathering, soil processes and 

21 preindustrial deposition, with units of equivalent charge/volume;

22 [ANClim] is the target ANC value, with units of equivalent charge/volume; 

 

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.
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1 Q is the catchment level runoff rate governed by water mass balance and dominated by

2 precipitation, with units of distance/time; and

3 Neco is the amount of nitrogen deposition that is effectively neutralized by a variety of biological 

4 (e.g., nutrient uptake) and physical processes, with units of equivalent charge/ (area-time).

5

6 Equation III- 1 is a modified expression that adopts the basic formulation of the steady

7 state models that are described in chapter 2 of the Policy Assessment. More detailed discussion 

8 of the rationale, assumptions and derivation of equation III- 1, as well as all of the equations in 

9 this section, are included in Appendix B of the Policy Assessment. The equation simply reflects 

10 the amount of deposition of nitrogen and sulfur from the atmosphere, CLANClim(N + S), that is 

11 associated with a sustainable long-term ANC target, [ANClim], given the capacity of the natural 

system to generate ANC, [BC]0

*

12 , and the capacity of the natural system to neutralize nitrogen 

13 deposition, Neco. This expression of critical load is valid when nitrogen deposition is greater 

than Neco.814 The runoff rate, Q, allows for balancing mass in the two environmental mediums –

15 atmosphere and catchment. This critical load expression can be focused on a single water system 

16 or more broadly. To extend applicability of the critical load expression (equation III-1) from the 

17 catchment level to broader spatial areas, the terms Qr and CLr, are used, which are the runoff rate 

18 and critical load, respectively, of the region over which all the atmospheric terms in the equation 

19 are defined. 

20 In considering the contributions of SOx or NOy species to acidification, it is useful to 

21 think of every depositing nitrogen atom as supplying one equivalent charge unit and every sulfur 

22 atom as depositing two charge units. The Policy Assessment uses equivalent charge per volume 

 

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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1 as a normalizing tool in place of the more familiar metrics such as mass or moles per volume. 

2 This allows for a clearer explanation of many of the relationships between atmospheric and 

3 ecosystem processes that incorporate mass and volume unit conventions somewhat specific to 

the environmental media of concern (e.g., m34 for air and liter for liquid water). Equivalent 

5 charge reflects the chemistry equilibrium fundamentals that assume electroneutrality, or 

6 balancing charge where the sum of cations always equals the sum of anions. 

7 As presented above, the terms S and N in the CLANClim (N + S) term broadly represent all 

8 species of sulfur or nitrogen that can contribute to acidifying deposition. This follows 

9 conventions used in the scientific literature that addresses critical loads, and it reflects all 

10 possible acidifying contributions from any sulfur or nitrogen species. For all practical purposes, 

11 S reflects SOx as described above, the sum of sulfur dioxide gas and particulate sulfate. 

12 However, N in equation III-1 includes both oxidized forms, consistent with the ambient 

13 indicator, NOy, in addition to the reduced nitrogen species, ammonia and ammonium ion, 

14 referred to as NHx. NHx is included in the critical load formulation because it contributes to 

15 potentially acidifying nitrogen deposition. Consequently, from a mass balance or modeling 

16 perspective, the form of the standard needs to account for NHx, as described below.

17 3. Linking Deposition to Ambient Air Indicators

18 The last major component of the form illustrated in Figure III-1 addresses the linkage 

19 between deposition of nitrogen and sulfur and concentrations of the ambient air indicators, NOy

20 and SOx. To link ambient air concentrations with deposition, the Policy Assessment defines a 

21 transference ratio, T, as the ratio of total wet and dry deposition to ambient concentration, 

22 consistent with the area and time period over which the standard is defined. To express 

23 deposition of NOy and SOx in terms of NOy and SOx ambient concentrations, two transference 
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1 ratios were defined, where TSOx equals the ratio of the combined dry and wet deposition of SOx 

2 to the ambient air concentration of SOx, and TNOy equals the ratio of the combined dry and wet 

3 deposition of NOy to the ambient air concentration of NOy.

4 As described in Chapter 7 of the Policy Assessment, reduced forms of nitrogen (NHX) are 

5 included in total nitrogen in the critical load equation, III-1. Reduced forms of nitrogen are 

6 treated separately, as are NOy and SOx, and the transference ratios are applied. This results in 

7 the following critical load expression that is defined explicitly in terms of the indicators NOy and 

8 SOx:

9

CLANClim(N + S) = ([BC]0

*

10 - [ANClim])Q + Neco = [NOy]TNOy + [SOx]TSOx + NHx (III-2)

11

12 This is the same equation as III-1, with the deposition associated with the critical load translated 

13 to deposition from ambient air concentrations via transference ratios. In addition, deposition of 

14 reduced nitrogen, oxidize nitrogen and oxidized sulfur are treated separately. 

15 Transference ratios are a modeled construct, and therefore cannot be compared directly to 

16 measurable quantities. There is an analogy to deposition velocity, as a transference ratio is 

17 basically an aggregated weighted average of the deposition velocities of all contributing species 

18 across dry and wet deposition, and transference ratio units are expressed as distance/time. 

19 However, wet deposition commonly is not interpreted as the product of a concentration times a 

20 velocity. Direct wet deposition observations are available which integrate all of the processes, 

21 regardless of how well they may be understood, related to wet deposition into a measurable 

22 quantity. There are reasonable analogies between the processes governing dry and wet 

23 deposition, from a fundamental mass transfer perspective. In both cases there is a transfer of 
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1 mass between the dry ambient phase and another medium, either a surface or vegetation in the 

2 case of dry deposition, or a rain droplet or cloud in the case of wet precipitation. The specific 

3 thermodynamic properties and chemical/biological reactions that govern the transfer of dry mass 

4 to plants or aqueous droplets differ, but either process can be based on conceptualizing the 

5 product of a concentration, or concentration difference, times a mass transfer coefficient which is 

6 analogous to the basic dry deposition model: dry deposition = concentration x velocity (US 

7 EPA, 2011, Appendix F). 

8 Transference ratios require estimates of wet deposition of NOy and SOx, dry deposition 

9 of NOy and SOx, and ambient air concentrations of NOy and SOx. Possible sources of 

10 information include model estimates or a combination of model estimates and observations, 

11 recognizing that dry deposition is a modeled quantity that can use observed or modeled estimates 

12 of concentration. The limited amount of NOy measurements in acid-sensitive areas as well as 

13 the combination of representative NOy, SO2 and SO4 observations generally preclude the use of 

14 observations for development of a standard that is applicable nationally.

15 The Policy Assessment considers a blending of observations and models to take 

16 advantage of their relative strengths; e.g., combining the NADP wet deposition observations, 

17 modeled dry deposition, and a mix of modeled and observed concentrations, using the model for 

18 those species not measured or measured with very sparse spatial coverage. A potential 

19 disadvantage of mixing and matching observations and model estimates is to lose consistency

20 afforded by using just modeling alone. A modeling platform like CMAQ is based on adhering to 

21 consistent treatment of mass conservation, by linking emission inputs with air concentrations and 

22 concentrations to deposition. Inconsistencies from combining processes from different analytical 

23 platforms increase the chance that mass (of nitrogen or sulfur) would unintentionally be 
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1 increased or decreased as the internal checking that assures mass conservation is lost. 

2 Transference ratios incorporate a broad suite of atmospheric processes and consequently 

3 an analytical approach that instills consistency in the linkage of these processes is preferable to 

4 an approach lacking such inherent consistency. This contention does not mean that observations 

5 alone, if available, could not be used, but suggests that the inconsistencies in combining models 

6 and observations for the purposes of developing transference ratios has the potential for creating 

7 unintended artifacts. 

8 While there is a reasonable conceptual basis for the concept of an aggregated deposition 

9 velocity referred to in the Policy Assessment as a transference ratio, there is very limited ability

10 to compare observed and calculated ratios. This is because the deposition velocity is dependent 

11 on individual species, and the mass transfer processes of wet and dry removal, while 

12 conceptually similar, are different. Consequently, there does not exist a meaningful approach to 

13 measure such an aggregated or lumped parameter. Therefore, at this time, the evaluation of 

14 transference ratios is based on sensitivity studies, analysis of variability, and comparisons with 

15 other models, as described in Appendix F of the Policy Assessment. 

16 As discussed in Appendix F, the interannual variability, as well as the sensitivity to 

17 emission changes of roughly 50%, results in changes of transference ratios of approximately 5 -

18 10%. Part of the reason for this inherent stability is due to the co-dependence of concentration 

19 and deposition. For example, as concentrations are reduced as a result of emissions reductions, 

20 deposition in turn is reduced since deposition is a direct linear function of concentration leading 

21 to negligible impact on the deposition-to-concentration ratio. Likewise, an overestimate of 

22 concentration likely does not induce a bias in the transference ratio. While it is important to 

23 continue to improve the model’s ability to match ambient concentrations in time and space, the 
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1 bias of a modeled estimate of concentration relative to observations does not necessarily result in 

2 a bias in a calculated transference ratio. In effect, this consideration of bias cancellation reduces 

3 the sensitivity of transference ratios to model uncertainties and affords increased confidence in 

4 the stability of these ratios. Based on the series of sensitivity and variability analyses, the Policy

5 Assessment concludes that the transference ratios are relatively stable and provide a sound 

6 metric for linking deposition and concentration in the form of the standard.

7 As discussed in the policy Assessment, transference ratios are dependent on the platform 

8 upon which they are constructed. Comparisons of transference ratios constructed from different 

9 modeling platforms do exhibit significant differences. While this divergence of results may be 

10 explained by a variety of differences in process treatments, input fields and incommensurabilities 

11 in species definitions and spatial configurations, it does suggest two very important conclusions. 

12 First, the idea of using multiple platforms for different parts of the country may be problematic 

13 as there does not exist a reliable approach to judge acceptance which is almost always based on 

14 comparisons to observations. Second, since transference ratios are based on concentrations and 

15 deposition, as the uncertainties in each of those components are reduced, the relative uncertainty 

16 in the ratios also is reduced. This means that basic improvements in the model’s ability to 

17 reproduce observed wet deposition and ambient concentration fields enhance the relative 

18 confidence in the constructed transference ratios. Similarly, as in-situ dry deposition flux 

19 measurements become available that enable a more rigorous evaluation and diagnosis of 

20 modeled dry deposition processes, the expected improved treatment of dry deposition also would 

21 increase confidence in transference ratios. Finally, deposition is directly related to ambient air 

22 concentrations. Models like CMAQ rely on the concentration-to-deposition linkage to calculate 

23 deposition, which is the foundation for broadly based and robust assessments addressing 
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1 atmospheric deposition. In principle, the use of a modeled constructed transference ratio is based 

2 on the same premise by which we use models to estimate deposition in the first place.

3 The shortage of widely available ambient air observations and the fact that estimates of 

4 dry deposition requires modeling, collectively suggests that a unified modeling platform is the 

5 best approach for constructing transference ratios. The Policy Assessment (US EPA, 

6 2011,section 2) considers CMAQ and other models, such as CAMx and the Canada’s AURAMS 

7 - A Unified Regional Air-quality Modeling System (Smythe et al., 2008), and concludes that 

8 CMAQ is the preferred modeling platform for constructing transference ratios for purposes of 

9 developing a new secondary standard for consideration. This conclusion reflects the view that 

10 for the purposes of defining transference ratios, a modeling platform should (1) be a multiple 

11 pollutant model recognizing the myriad of connections across pollutant categories that directly

12 and indirectly impact nitrogen and sulfur characterization, (2) include the most comprehensive 

13 scientific treatments of atmospheric processes that relate directly and indirectly to characterizing 

14 concentrations and deposition, (3) have an infrastructure capability that accommodates the 

15 inclusion of improved scientific treatments of relevant processes and important input fields, and 

16 (4) undergo frequent reviews regarding the adequacy of the underlying science as well as the 

17 appropriateness in applications. The CMAQ platform exhibits all these characteristics. It has 

18 been (and continues to be) extensively evaluated for several pollutant categories, is supported by

19 a central infrastructure of EPA scientists, with considerable interfacing with the scientific 

20 research communities in academia and industry, whose mission is to improve and evaluate the 

21 CMAQ platform. More directly, CMAQ, and its predecessor versions, has a long track record 

22 going back to the NAPAP in the 1980’s of specific improvements in deposition processes, which 

23 are described in Appendix F of the Policy Assessment.
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1 4. Aquatic Acidification Index

2 Having established the various expressions that link atmospheric deposition of nitrogen

3 and sulfur to ANC and the transference ratios that translate atmospheric concentrations to 

4 deposition of nitrogen and sulfur, the Policy Assessment derives the following expression of 

5 these linkages, which separates reduced forms of nitrogen, NHx, from oxidized forms:

6

7 ANCcalc = {ANClim + CLr/Qr } – NHx/Qr - TNOy [NOy]/Qr - TSOx[SOx]/Qr (III-3)

8

9 Equation III- 3 is the basic expression of the form of a standard that translates the 

10 conceptual framework into an explicit expression that defines ANC as a function of the ambient 

11 air indicators, NOy and SOx. This equation calculates an expected ANC value based on ambient 

12 concentrations of NOy and SOx. The calculated ANC will differ from the target ANC (ANClim) 

13 depending on how much the nitrogen and sulfur deposition associated with NOy, SOx, and NHx 

14 differs from the critical load associated with just achieving the target ANC. If the sum of the 

15 three deposition terms is less than or equal to the critical load associated with the target ANC 

16 level (ANClim), the standard is achieved.

17 Based on equation III- 3, the Policy Assessment defines an aquatic acidification index (AAI) that 

18 is more simply stated using terms that highlight the ambient air indicators:

19

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

21

22 where the AAI represents the long term (or steady state) ANC level associated with ambient air 

23 concentrations of NOy and SOx. The factors F1 through F4 convey three attributes: a relative 
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1 measure of the ecosystem’s ability to neutralize acids (F1), the acidifying potential of reduced 

2 nitrogen deposition (F2), and the deposition-to-concentration translators for NOy (F3) and SOx 

3 (F4). 

4 Specifically:

5 F1 = ANClim + CLr/Qr ;

6 F2 = NHx/Qr = NHx deposition divided by Qr;

7 F3 = TNOy/ Qr ; TNOy is the transference ratio that converts ambient air concentrations of NOy to 

8 deposition of NOy; and

9 F4 = TSOx/ Qr ; TSOx is the transference ratio that converts ambient air concentrations of SOx to 

10 deposition of SOx.

11 All of these factors include representative Qr to maintain unit (and mass) consistency between 

12 the AAI and the terms on the right side of equation III-4.

13 The F1 factor is the target ANC level plus the amount of deposition (critical load) the 

14 ecosystem can receive and still achieve the target level. It incorporates an ecosystem’s ability to 

15 generate acid neutralizing capacity through base cation supply ([BC]*0) and to neutralize 

16 acidifying nitrogen deposition through Neco, both of which are incorporated in the CL term. As 

17 noted above, because Neco can only neutralize nitrogen deposition (oxidized or reduced) there 

18 may be rare cases where Neco exceeds the combination of reduced and oxidized nitrogen 

19 deposition. Consequently, to ensure that the AAI equation is applicable in all cases that may

20 occur, equation III-4 is conditional on total nitrogen deposition, {NHx + F3[NOy]}, being 

21 greater than Neco. In rare cases where Neco is greater than {NHx + F3[NOy]}, F2, F3, and 

22 Neco would be set equal to 0 in the AAI equation. The consequence of setting F2 and F3 to zero 
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1 is simply to constrain the AAI calculation just to SOx, as nitrogen would have no bearing on 

2 acidifying contributions in this case.

3 The Policy Assessment concludes that equation III- 4 (US EPA, 2011,equation 7-12),

4 which defines the AAI, is ecologically relevant and appropriate for use as the form of a national 

5 standard designed to provide protection for aquatic ecosystems from the effects association with 

6 acidifying deposition associated with concentrations of oxides of nitrogen and sulfur in the 

7 ambient air. This AAI equation does not, however, in itself, define the spatial areas over which 

8 the terms of the equation would apply. To specify values for factors F1 through F4, it is 

9 necessary to define spatial areas over which these factors are determined. Thus, it is necessary to 

10 identify an approach for spatially aggregating water bodies into ecologically meaningful regions 

11 across the U.S., as discussed below.

12 5. Spatial Aggregation

13 As discussed in the Policy Assessment, one of the unique aspects of this form is the need 

14 to consider the spatial areas over which values for the F factors in the AAI equation are 

15 quantified. Ecosystems across the U.S. exhibit a wide range of geological, hydrological and 

vegetation characteristics that influence greatly the ecosystem parameters, Q, BC0

*

16 and Neco that 

17 are incorporated in the AAI. Variations in ecosystem attributes naturally lead to wide variability 

18 in the sensitivities of water bodies in the U.S. to acidification, as well as in the responsiveness of 

19 water bodies to changes in acidifying deposition. Consequently, variations in ecosystem 

20 sensitivity, and the uncertainties inherent in characterizing these variations, must be taken into 

21 account in developing a national standard. In developing a secondary NAAQS to protect public 

22 welfare, the focus of the Policy Assessment is on protecting sensitive populations of water 

23 bodies, not on each individual water body, which is consistent with the Agency’s approach to 
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1 protecting public health through primary NAAQS that focus on susceptible populations, not on 

2 each individual.

3 The approach used for defining ecologically relevant regions across the U.S. in the Policy 

4 Assessment (US EPA, 2011,section 7.2.5) is described below, along with approaches to 

5 characterizing each region as acid sensitive or relatively non-acid sensitive. This 

6 characterization facilitates a more detailed analysis and focus on those regions that are relatively 

7 more acid sensitive. This characterization is also used to avoid over-protection in relatively non8 acid sensitive regions, regions that would receive limited benefit from reductions in the 

9 deposition of oxides of nitrogen and sulfur with respect to aquatic acidification effects. 

10 Approaches to developing representative values for each of the terms in the AAI equation 

11 (factors F1 through F4) for each ecologically relevant region for which sufficient data are 

12 available are then discussed. These spatial aggregation approaches are generally applicable to 

13 the contiguous United States. The following discussion also addresses the development of 

14 factors for data-limited regions and specifically for Hawaii, Alaska and the U.S. territories.

15 Stated more simply, this section discusses appropriate ways to divide the country into 

16 ecologically relevant regions; to characterize each region as either acid sensitive or relatively 

17 non-acid sensitive; and to determine values of factors F1 through F4 for each region, taking into 

18 consideration the acid sensitivity of each region and the availability of relevant data. For each 

19 such region, the AAI would be calculated based on the values of factors F1 through F4 specified 

20 for that region.

21 In considering approaches to spatial aggregation, the Policy Assessment focuses on 

22 methods that have been developed to define ecologically relevant regions, referred to as 

23 ecoregions, which are meaningfully related to the factors that are relevant to aquatic 
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1 acidification. As noted above, the Policy Assessment did not focus on looking at each individual 

2 water body, nor did it focus on aggregating over the entire nation, which would preclude taking 

3 into account the inherent variability in atmospheric and ecological factors that fundamentally 

4 modify the relationships that are central to the development of an ecologically relevant AAI. 

5 Based on considering available classification schemes, the Policy Assessment concludes 

6 that Omernik’s ecoregion classification is the most appropriate method to consider for the 

7 purposes of this review. This classification offers several levels of spatial delineation, has 

8 undergone an extensive scientific peer review process, and has explicitly been applied to 

9 delineating acid sensitive areas within the U.S. Further, the Policy Assessment concludes that 

ecoregion level III (Figure III-1) resolution, with 84 defined ecoregions in the contiguous U.S.,910

11 is the most appropriate level to consider for this purpose. The spatial resolution afforded by 

12 level III strikes an appropriate balance relative to the reasoning that supports conclusions on 

13 indicators, as discussed above. The Policy Assessment concludes that the most detailed level of 

14 resolution (level IV) is not appropriate given the limited data availability to address nearly 1000 

15 subdivisons within that level and the currently evolving nature of level IV regions. Further, level 

16 III ecoregions are preferred to level II in that level III ecoregions, but not level II ecoregions, are 

17 largely contiguous in space which allows for a more coherent development of information to 

18 quantify the AAI factors and to characterize the concentrations of NOy and SOx in the ambient 

19 air within each ecoregion.

20 Appendix C of the Policy Assessment includes a description of each level III ecoregion. 

21 The Policy Assessment notes that the use of ecoregions is an appropriate spatial aggregation 

22 scheme for this standard focused on deposition-related aquatic acidification effects, while many 

23 of the same ecoregion attributes may be applicable in subsequent NAAQS reviews that may

 

9 We note that an 85th area within Omernik’s Ecoregion Level III is currently being developed for California.
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1 address other deposition-related aquatic and terrestrial ecological effects. Because atmospheric 

2 deposition is modified by ecosystem attributes, the types of vegetation, soils, bedrock geology, 

3 and topographic features that are the basis of this ecoregion classification approach also will 

4 likely be key attributes for other deposition-related effects (e.g., terrestrial acidification, nutrient 

5 enrichment) that link atmospheric concentrations to an aquatic or terrestrial ecological indicator.

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

7

8 a. Ecoregion Sensitivity

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

10 2) and more recent ANC data to delineate two broad groupings of ecoregions: acid-sensitive and 
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1 relatively non-acid sensitive ecoregions. This delineation was made to facilitate greater focus on 

2 those ecoregions with water bodies that generally have greater acid sensitivity and to avoid over3 protection in regions with generally less sensitive water bodies. The approach used to delineate 

4 acid-sensitive and relatively non-acid sensitive regions included an initial numerical-based 

5 sorting scheme using ANC data, which categorized ecoregions with relatively high ANC values 

6 as being relatively non-acid sensitive. This initial delineation resulted in 29 of the 84 Omernik 

7 ecoregions being categorized as acid sensitive. Subsequently, land use data were also considered 

8 to determine to what extent an ecoregion is of a relatively pristine and rural nature by quantifying 

9 the degree to which active management practices related to development and agriculture occur in 

10 each ecoregion. 

11 The overall objective is to produce a logical and practical grouping of ecoregions that 

12 experience adverse conditions with respect to aquatic acidification and are likely to respond to 

13 changes in concentrations of NOy and SOx in the ambient air and to the related deposition 

14 levels. To achieve this goal, a two-step process has been applied, first identifying acid sensitive 

15 ecoregions based on water quality data alone, and second identifying among those acid-sensitive 

16 ecoregions those with highly developed and managed areas. These highly developed and 

17 managed ecoregions are placed in a non-acid sensitive category to avoid over protection beyond 

18 what is requisite to protect public welfare. More specifically, in determining an ecoregion’s acid 

19 sensitivity status in step 1, ANC data across the 84 ecoregions are sorted (US EPA, 2011,section

20 7 ) to determine the number of water bodies within a region with ANC values suggestive of acid 

21 sensitivity, so as to screen out regions with an overabundance of high ANC values. In reviewing 

22 the ANC data, the Policy Assessment identified 29 ecoregions that meet two criteria: (1) greater 

23 than 5% of water bodies with data with ANC values less than 200 μeq/L and (2) greater than 1% 
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1 of water bodies with ANC values less than 100 μeq/L. In step 2, land use data were used to 

2 identify those acid sensitive ecoregions with significant managed areas that would not be 

3 considered as having a relatively pristine and rural character. The percentage of the 

4 combination of developed (residential, transportation, industrial and commercial) and 

5 agricultural (croplands, pastures, orchards, vineyards) land use was used as an indicator of 

6 managed land use area. Forest cover was used as an indicator of non-managed land use more 

7 directly reflecting the pristine quality of a region. Based on the 2006 National Land Cover Data 

8 base (NLCD, http://www.epa.gov/mrlc/nlcd-2006.html), acid sensitive ecoregions would meet 

9 both of the following land use data criteria: percent of developed and agricultural area less than 

10 20% combined with forested area greater than 50%. The combination of steps 1 and 2 identify

11 22 relatively acid sensitive areas (Table III-1 and Figure III-2). 

12 Consideration was also given to the use of naturally acidic conditions in defining 

13 relatively non-acid sensitive areas. For example, several of the ecoregions located in plains near 

14 the coast exhibit elevated dissolved organic carbon (DOC) levels, which is associated with 

15 naturally acidic conditions. DOC in surface waters is derived from a variety of weak organic 

16 acid compounds generated from the natural availability and decomposition of organic matter 

17 from biota. Consequently, high DOC is associated with “natural” acidity, with the implication 

18 that a standard intended to protect against atmospheric contributions to acidity is not an area of 

19 focus. The evidence suggests that several of the more highly managed ecoregions in coastal or 

20 near coastal transition zones are associated with relatively high DOC values, typically exceeding 

21 on average 5 mg/l, compared to other acid sensitive areas. Although there is sound logic to 

22 interpret naturally acidic areas as relatively non-acid sensitive, natural acidity indicators were not 

23 explicitly included in defining relatively non-acid sensitive areas as there does not exist a 
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1 consensus-based quantifiable scientific definition of natural acidity. Approaches to explicitly

2 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

3
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1

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

3

4 b. Representative Ecoregion-specific Factors

5 Having concluded that the Omernik level III ecoregions are an appropriate approach to 

6 spatial aggregation for the purpose of a standard to address deposition-related aquatic 

7 acidification effects, the Policy Assessment uses those ecoregions to define each of the factors in 

8 the AAI equation. As discussed below, factors F1 through F4 in equation III-4 are defined for 

9 each ecoregion by specifying ecoregion-specific values for each factor based on monitored or 

10 modeled data that are representative of each ecoregion.
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1 i. Factor F1

2 As discussed above, factor F1 reflects a relative measure of an ecosystem’s ability to 

3 neutralize acidifying deposition, and is defined as: F1 = ANClim + CLr/Qr. The value of F1 for 

4 each ecoregion would be based on a representative critical load for the ecoregion (CLr) 

5 associated with a single national target ANC level (ANClim, discussed below in section III.D), 

6 as well as on a representative runoff rate (Qr). To specify ecoregion-specific values for the term 

Qr 7 , the Policy Assessment used the median value of the distribution of Q values that are available 

8 for water bodies within each ecoregion. To specify ecoregion-specific representative values for 

the term CLr in factor F1, a distribution10 9 of calculated critical loads was created for the water 

bodies in each ecoregion for which sufficient water quality and hydrology data are available.11 10 

11 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 12 percentile means that within an 

13 ecoregion, 90 percent of the water bodies would be expected to have higher calculated critical 

14 loads than the representative critical load. That is, if the representative critical load were to 

15 occur across the ecoregion, 90 percent of the water bodies would be expected to achieve the 

16 national ANC target or better.

17 The specific percentile selected as part of the definition of F1 is an important parameter 

18 that directly impacts the representative critical load specified for each ecoregion, and therefore 

19 the degree of protectiveness of the standard. A higher percentile corresponds to a lower critical 

20 load and, therefore, to lower allowable ambient air concentrations of NOy and SOx and related 

 

10 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.

11 The Policy Assessment judged the data to be sufficient for this purpose if data are available from more than 10 

water bodies in an ecoregion.
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1 deposition to achieve a target AAI level. In conjunction with the other terms in the AAI

2 equation, alternative forms can be appropriately characterized in part by identifying a range of 

3 alternative percentiles. The choice of an appropriate range of percentiles to consider for acid4 sensitive and relatively non-acid sensitive ecoregions, respectively, is discussed below. 

5 (a) Acid-sensitive Ecoregions

6 In identifying percentiles that are appropriate to consider for the purpose of calculating 

7 factor F1 for ecoregions characterized as acid sensitive, the Policy Assessment concludes that it 

8 is appropriate to focus on the upper part of the distribution of critical loads, so as to ensure that 

9 the ecoregion would be represented by relatively more acid sensitive water bodies within the 

10 ecoregion. Specifying factor F1 in this way would help to define a standard that would be 

11 protective of the population of acid sensitive water bodies within an ecoregion, recognizing that 

12 even ecoregions characterized as acid sensitive may contain a number of individual water bodies 

13 that are not acid sensitive. The Policy Assessment recognizes that there is no basis for 

14 independently evaluating the degree of protectiveness afforded by any specific percentile value, 

15 since it is the combination of form and level, in conjunction with the indicator and averaging 

16 time, which determine the degree of protectiveness of a standard. In light of this, the Policy

17 Assessment concludes that it is appropriate to consider a range of percentiles, from well above 

the 50th 18 percentile, or median, of the distribution to somewhat below the highest value. More 

19 specifically, the Policy Assessment concludes it is appropriate to consider percentiles in the 

range of the 70th to the 90th 20 percentile. This conclusion is based on the judgment that it would 

21 not be appropriate to represent an ecoregion with the lowest or near lowest critical load, so as to 

22 avoid potential extreme outliers that can be seen to exist at the extreme end of the data

23 distributions, which would not be representative of the population of acid sensitive water bodies 
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1 within the ecoregion and could lead to an overly protective standard. At the same time, in 

2 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 3 percentile, a value well above the median 

4 of the distribution, so that a substantial majority of acid-sensitive water bodies are protected. 

5 In considering this conclusion, the CASAC Panel noted that the data bases for calculating 

6 critical loads within an ecoregion are not necessarily representative of all water bodies within an 

7 ecoregion. That is, in many ecoregions the lake sampling design used in studies that generated 

8 the relevant data may have focused on the relatively more sensitive water bodies within an 

9 ecoregion (Russell and Samet, 2011a). Consequently, a given percentile of the distribution of 

10 calculated critical loads, based on sampled water bodies, may not be representative of that 

11 percentile of all water bodies across an entire ecoregion. To the extent that the sampling of water 

12 bodies within an ecoregion was skewed toward the relatively more sensitive water bodies, 

13 selecting a given percentile from the distribution of available critical loads would result in a 

14 somewhat higher percentile of all water bodies within that ecoregion having a higher calculated 

15 critical load than the representative critical load value. Thus, the extent to which study sampling 

16 designs have resulted in skewed distributions of calculated critical loads is an uncertainty that is 

17 appropriate to consider in selecting a percentile for the purpose of defining the factor F1 in the 

18 AAI equation. 

19 (b) Non-acid sensitive Ecoregions

20 With regard to identifying percentiles that are appropriate to consider for the purpose of 

21 calculating factor F1 for ecoregions characterized as relatively non-acid sensitive, the Policy

22 Assessment recognizes that while such ecoregions are generally less sensitive to acidifying 

23 deposition from oxides of nitrogen and sulfur, they may contain a number of water bodies that 
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1 are acid sensitive This category includes ecoregions that are well protected from acidification 

2 effects due to natural production of base cations and high ANC levels, as well as naturally acidic 

3 systems with limited base cation production and consequently very low critical loads. Therefore, 

4 the use of a critical load that would be associated with a highly sensitive water body in a 

5 naturally acidic system would impose a high degree of relative protection in terms of allowable 

6 ambient air concentrations of oxides of nitrogen and sulfur and related deposition, while 

7 potentially affording little or no public welfare benefit from attempting to improve a naturally

8 acidic system.

9 Based on these considerations, the Policy Assessment concludes it is appropriate to 

10 consider the use of a range of percentiles that extends lower than the range identified above for 

11 acid-sensitive ecoregions. Consideration of a lower percentile would avoid representing a 

12 relatively non-acid sensitive ecoregion by a critical load associated with relatively more acid13 sensitive water bodies. In particular, the Policy Assessment concludes it is appropriate to focus 

on the median or 50th 14 percentile of the distribution of critical loads so as to avoid over-protection 

15 in such ecoregions. Recognizing that relatively non-acid sensitive ecoregions generally are not 

16 sampled to the extent that acid-sensitive ecoregions are, it also is appropriate to consider using 

17 the median critical load of all relatively non-acid sensitive ecoregions for each such ecoregion.

18 ii. Factor F2

19 As discussed above, factor F2 is the amount of reduced nitrogen deposition within an 

20 ecoregion, including the deposition of both ammonia gas and ammonium ion, and is defined as: 

F2 = NHx/Qr. The Policy Assessment calculated the representative runoff rate, Qr 21 , using a 

22 similar approach as noted above for factor F1; i.e., the median value of the distribution of Q 

23 values that are available for water bodies within each ecoregion. In the Policy Assessment, 2005 
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1 CMAQ model simulations over 12-km grids are used to calculate an average value of NHx for 

2 each ecoregion. The NHx term is based on annual average model outputs for each grid cell, 

3 which are spatially averaged across all the grid cells contained in each ecoregion to calculate a 

4 representative annual average value for each ecoregion. The Policy Assessment concludes that 

5 this approach of using spatially averaged values is appropriate for modeling, largely due to the 

6 relatively rapid mixing of air masses that typically results in relatively homogeneous air quality 

7 patterns for regionally dispersed pollutants. In addition, there is greater confidence in using 

8 spatially averaged modeled atmospheric fields than in using modeled point-specific fields.

9 This averaging approach is also used for the air concentration and deposition terms in 

10 factors F3 and F4, as discussed below. The Policy Assessment notes that modeled NHx 

11 deposition exhibits greater spatial variability than the other modeled terms in factors F3 and F4. 

12 Recognizing this greater variability, the Policy Assessment concludes that it would be 

13 appropriate to consider alternative approaches to specifying the value of NHx. One such 

14 approach might involve the use of more localized and/or contemporaneous modeling in areas 

15 where this term is likely to be particularly variable and important.

16 iii. Factors F3 and F4

17 As discussed above, factors F3 and F4 are the ratios that relate ambient air concentrations 

18 of NOy and SOx to the associated deposition, and are defined as follow: F3 = TNOy/ Qr and F4 

19 = TSOx/ Qr. TNOy is the transference ratio that converts ambient air concentrations of NOy to 

20 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 21 , is calculated as for factors F1 and 

22 F2. The transference ratios are based on the 2005 CMAQ simulations, using average values for 

23 each ecoregion, as noted above for factor F2. More specifically, the transference ratios are 
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1 calculated as the annual deposition of NOy or SOx spatially averaged across the ecoregion and 

2 divided by the annual ambient air concentration of NOy or SOx, respectively, spatially averaged 

3 across the ecoregion.

4 c. Factors in Data-limited Ecoregions 

5 As discussed above in section III.B.5.a, in the Policy Assessment the initial delineation of 

6 acid-sensitive and relatively non-acid sensitive ecoregions was based on available ANC and 

7 alkalinity data. Areas not meeting the ANC criteria described above are categorized as 

8 relatively non-acid sensitive. The development of a reasonable distribution of critical loads for 

9 water bodies within an ecoregion for the purpose of identifying the representative critical load 

10 requires additional data, including more specific water quality data for major cations and anions. 

11 This means that the water bodies that can be used to develop a distribution of critical loads is 

12 generally a subset of those water bodies for which ANC data are available Consequently, there 

13 are certain ecoregions with sparse data that are not suitable for developing a distribution of 

14 critical loads.

15 As noted above, the Policy Assessment judges that it is not appropriate to develop such 

16 distributions based on data from less than ten water bodies within an ecoregion. Twelve such 

17 ecoregions, which included only relatively non-acid sensitive ecoregions, were characterized as 

18 being data-limited. For these ecoregions, the Policy Assessment considered alternative 

19 approaches to specifying values for the terms CLr and Qr for the purpose of determining values 

20 for each of the factors in the AAI equation. For these data-limited ecoregions, the Policy

21 Assessment judges that it is appropriate to use the median values of CLr and Qr from the 

22 distributions of these terms for all other relatively non-acid sensitive ecoregions, rather than 

23 attempting to use severely limited data to develop a value for these terms based solely on data 
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1 from such an ecoregion. Further, consideration could be given to using a single national default 

2 value for all relatively non-acid sensitive ecoregions. The Policy Assessment notes that this data 

3 limitation is not a concern in specifying values for the other terms in the AAI equation for such 

4 ecoregions, since those terms are based on data from the 2005 CMAQ model simulation, which 

5 covers all ecoregions across the contiguous U.S.

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

7 The above methods for specifying ecoregion-specific values for the factors in the AAI 

8 equation apply to those ecoregions within the contiguous U.S. For areas outside the continental 

9 U.S., including Hawaii, Alaska, and the U.S. Territories, there is currently a lack of available 

10 data to characterize the sensitivity of such areas, as well as a lack of water body-specific data and 

11 CMAQ-type modeling to specify values for the F1 through F4 factors. Thus, the Policy

12 Assessment has considered possible alternative approaches to specifying values for factors F1 

13 through F4 in the AAI equation for these areas.

14 One such approach could be to specify area-specific values for the factors based on 

15 values derived for ecoregions with similar acid sensitivities, to the extent that relevant 

16 information can be obtained to determine such similarities. Such an approach would involve 

17 conducting an analysis to characterize similarities in relevant ecological attributes between 

18 ecoregions in the contiguous U.S. and these areas outside the contiguous U.S. so as to determine 

19 the appropriateness of utilizing ecoregion-specific values for the CLr and Qr terms from one or 

20 more ecoregions within the contiguous U.S. This approach would also involve conducting 

21 additional air quality modeling for these area that are outside the geographical scope of the 

22 currently available CMAQ model simulations, so as to develop the other information necessary 

23 to specify values for factors F2 through F4 for these areas.
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1 A second approach could rely on future data collection efforts to establish relevant 

2 ecological data within these areas that, together with additional air quality modeling, could be 

3 used to specify area-specific values for factors F1 through F4. Until such time as relevant data 

4 become available, these areas could be treated the same as data-limited ecoregions in the 

5 contiguous U.S. that are relatively non-acid sensitive.

6 The Policy Assessment concludes that either approach would introduce substantial 

7 uncertainties that arise from attempting to extrapolate values based on similarity assumptions or 

8 arbitrarily assigning values for factors in the AAI equation that would be applicable to these 

9 areas outside the contiguous U.S. In light of such uncertainties, the Policy Assessment 

10 concludes that it would also be appropriate to consider relying on the existing NO2 and SO2

11 secondary standards in these areas for protection of any potential direct or deposition-related 

12 ecological effects that may be associated with the presence of oxides of nitrogen and sulfur in the 

13 ambient air. The Policy Assessment concludes that relying on existing secondary standards in 

14 these areas is preferable to using a highly uncertain approach to allow for the application of a 

15 new standard based on the AAI in the absence of relevant area-specific data.

16 6. Summary of the AAI Form

17 With regard to the form of a multi-pollutant air quality standard to address deposition18 related aquatic acidification effects, the Policy Assessment concludes that consideration should 

19 be given to an ecologically relevant form that characterizes the relationships between the 

20 ambient air indicators for oxides of nitrogen and sulfur, the related deposition of nitrogen and 

21 sulfur, and the associated aquatic acidification effects in terms of a relevant ecological indicator. 

22 Based on the available information and assessments, consideration should be given to using acid 

23 neutralizing capacity (ANC) as the most appropriate ecological indicator for this purpose, in that 
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1 it provides the most stable metric that is highly associated with the water quality properties that 

2 are directly responsible for the principal adverse effects associated with aquatic acidification: 

3 fish mortality and reduced aquatic species diversity.

4 The Policy Assessment developed such a form, termed an aquatic acidification index 

5 (AAI), using a simple equation to calculate an AAI value in terms of the ambient air indicators of 

6 oxides and nitrogen and sulfur and the relevant ecological and atmospheric factors that modify

7 the relationships between the ambient air indicators and ANC. Recognizing the spatial 

8 variability of such factors across the U.S., the Policy Assessment concludes it is appropriate to 

9 divide the country into ecologically relevant regions, characterized as acid-sensitive or relatively 

10 non-acid-sensitive, and specify the value of each of the factors in the AAI equation for each such 

11 region. Omernik ecoregions, level III, are identified as the appropriate set of regions over which 

12 to define the AAI. There are 84 such ecoregions that cover the continental U.S. This set of 

13 ecoregions is based on grouping a variety of vegetation, geological, and hydrological attributes 

14 that are directly relevant to aquatic acidification assessments and that allow for a practical 

15 application of an aquatic acidification standard on a national scale.

16 The Policy Assessment defines AAI by the following equation: AAI = F1 – F2 –

17 F3[NOy] – F4[SOx]. Factors F1 through F4 would be defined for each ecoregion by specifying 

18 ecoregion-specific values for each factor based on monitored or modeled data that are 

19 representative of each ecoregion. The F1 factor is also defined by a target ANC value. More 

20 specifically:

21 (1) F1 reflects a relative measure of an ecosystem’s ability to neutralize acidifying 

22 deposition. The value of F1 for each ecoregion would be based on a representative 

23 critical load for the ecoregion associated with a single national target ANC level, as well 
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1 as on a representative runoff rate. The representative runoff rate, which is also used in 

2 specifying values for the other factors, would be the median value of the distributions of 

3 runoff rates within the ecoregion. The representative critical load would be derived from 

4 a distribution of critical loads calculated for each water body in the ecoregion for which 

5 sufficient water quality and hydrology data are available. The representative critical load 

6 would be defined by selecting a specific percentile of the distribution.

7 In identifying a range of percentiles that are appropriate to consider for this 

8 purpose, regions categorized as acid sensitive were considered separately from regions 

9 categorized as relatively non-acid sensitive. For acid sensitive regions, the Policy

10 Assessment concludes that consideration should be given to selecting a percentile from 

within the range of the 70th to the 90th 11 percentile. The lower end of this range was 

12 selected to be appreciably above the median value so as to ensure that the critical load 

13 would be representative of the population of relatively more acid sensitive water bodies 

14 within the region, while the upper end was selected to avoid the use of a critical load 

15 from the extreme tail of the distribution which is subject to a high degree of variability

16 and potential outliers. For relatively non-acid sensitive regions, the Policy Assessment 

concludes that consideration should be given to selecting the 50th 17 percentile to best 

18 represent the distribution of water bodies within such a region, or alternatively to using 

19 the median critical load of all relatively non-acid sensitive areas, recognizing that such 

20 areas are far less frequently evaluated than acid sensitive areas. Using either of these 

21 approaches would avoid characterizing a generally non-acid-sensitive region with a 

22 critical load that is representative of relatively acid sensitive water bodies that may exist 

23 within a generally non-acid sensitive region.
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1 (2) F2 reflects the deposition of reduced nitrogen. Consideration should be given to 

2 specifying the value of F2 for each region based on the averaged modeled value across 

3 the region, using national CMAQ modeling that has been conducted by EPA. 

4 Consideration could also be given to alternative approaches to specifying this value, such 

5 as the use of more localized and/or contemporaneous modeling in areas where this term is 

6 likely to be particularly variable and important. 

7 (3) F3 and F4 reflect transference ratios that convert ambient air concentrations of NOy

8 and SOx, respectively, into related deposition of nitrogen and sulfur. Consideration 

9 should be given to specifying the values for F3 and F4 for each region based on CMAQ 

10 modeling results averaged across the region. We conclude that specifying the values or 

11 the transference rations based on CMAQ modeling results alone is preferred to an 

12 alternative approach that combines CMAQ model estimates with observational data.

13 (4) The terms [NOy] and [SOx] reflect ambient air concentrations measured at 

14 monitoring sites within each region.

15 Using the equation, a value of AAI can be calculated for any measured values of ambient 

16 NOy and SOx. For such a NAAQS, the Administrator would set a single, national value for the 

17 level of the AAI used to determine achievement of the NAAQS, as discussed below in section 

18 III.D. The ecoregion-specific values for factors F1 through F4 would be specified by EPA based 

19 on the most recent data and CMAQ model simulations, and codified as part of such a standard. 

20 These factors would be reviewed and updated as appropriate in the context of each periodic 

21 review of the NAAQS. 

22 The Policy Assessment developed specific F factors for each ecoregion based on the 

23 approach discussed above, using alternative percentiles and alternative national target ANC 
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1 levels. The results of this analysis for ecoregions characterized as acid sensitive are presented in 

2 Table 7-1a-d in the Policy Assessment. 

3 C. Averaging Time

4 As discussed in section 7.3 of the Policy Assessment, aquatic acidification can occur over 

5 both long- and short-term timescales. Long-term cumulative deposition of nitrogen and sulfur is 

6 reflected in the chronic acid-base balance of surface waters as indicated by measured annual 

7 ANC levels. Similarly, the use of steady state critical load modeling, which generates critical 

8 loads in terms of annual cumulative deposition of nitrogen and sulfur, means that the focus of 

9 ecological effects studies based on critical loads is on the long-term equilibrium status of water 

10 quality in aquatic ecosystems. Much of the evidence of adverse ecological effects associated 

11 with aquatic acidification, as discussed above in section II.A, is associated with chronically low 

12 ANC levels. Protection against a chronic ANC level that is too low is provided by reducing 

13 overall annual average deposition levels for nitrogen and sulfur.

14 Reflecting this focus on long-term acidifying deposition, the Policy Assessment 

15 developed the AAI that links ambient air indicators to deposition-related ecological effects, in 

16 terms of several factors, F1 through F4. As discussed above, these factors are all calculated as 

17 annual average values, whether based on water quality and hydrology data or on CMAQ model 

18 simulations. In the context of a standard defined in terms of the AAI, the Policy Assessment 

19 concludes that it is appropriate to consider the same annual averaging time for the ambient air 

20 indicators as is used for the factors in the AAI equation.

21 We also recognize that short-term (i.e., hours or days) episodic changes in water 

22 chemistry, often due to changes in the hydrologic flow paths, can have important biological 

23 effects in aquatic ecosystems. Such short-term changes in water chemistry are termed “episodic 
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1 acidification.” Some streams may have chronic or base flow chemistry that is generally healthy 

2 for aquatic biota, but may be subject to occasional acidic episodes with potentially lethal 

3 consequences. Thus, short-term episodic ecological effects can occur even in the absence of 

4 long-term chronic acidification effects.

5 Episodic declines in pH and ANC are nearly ubiquitous in drainage waters throughout the 

6 eastern U.S. Episodic acidification can result from several mechanisms related to changes in 

7 hydrologic flow paths. For example, snow can store nitrogen deposited throughout the winter 

8 and snowmelt can then release this stored nitrogen, together with nitrogen derived from 

9 nitrification in the soil itself, in a pulse that leads to episodic acidification in the absence of 

10 increased deposition during the actual episodic acidification event. The Policy Assessment notes 

11 that inputs of nitrogen and sulfur from snowpack and atmospheric deposition largely cycle 

12 through soil. As a result, short-term direct deposition inputs are not necessarily important in 

13 episodic acidification. Thus, as noted in chapter 3 of the ISA, protection against episodic acidity

14 events can be achieved by establishing a higher chronic ANC level.

15 Taken together, the above considerations support the conclusion that it is appropriate to 

16 consider the use of a long-term average for the ambient air indicators NOy and SOx for an 

17 aquatic acidification standard defined in terms of the AAI. The use of an annual averaging time 

18 for NOy and SOx concentrations would be appropriate to provide protection against low chronic 

19 ANC levels, which in turn would protect against both long-term acidification and acute acidic 

20 episodes.

21 The Policy Assessment has also considered interannual variability in both ambient air 

22 quality and in precipitation, which is directly related to the deposition of oxides of nitrogen and 

23 sulfur from the ambient air. While ambient air concentrations show year-to-year variability, 
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1 often the year-to-year variability in precipitation is considerably greater, given the highly 

2 stochastic nature of precipitation. The use of multiple years over which annual averages are 

3 determined would dampen the effects of interannual variability in both air quality and 

4 precipitation. For the ambient air indicators, the use of multiple-year averages would also add 

5 stability to calculations used to judge whether an area meets a standard defined in terms of the 

6 AAI. Consequently, the Policy Assessment concludes that an annual averaging time based on 

7 the average of each year over a consecutive 3 to 5 year period is appropriate to consider for the 

8 ambient air indicators NOy and SOx. In reaching this conclusion, the Policy Assessment notes 

9 that in its comments on the second draft Policy Assessment, CASAC agreed that a 3 to 5 year 

10 averaging time was appropriate to consider (Russell and Samet, 2010b).

11 D. Level

12 As discussed above, the Policy Assessment concludes that ANC is the ecological 

13 indicator best suited to reflect the sensitivity of aquatic ecosystems to acidifying deposition from 

14 oxides of nitrogen and sulfur in the ambient air. ANC is an indicator of the aquatic acidification 

15 expected to occur given the natural buffering capacity of an ecosystem and the loadings of 

16 nitrogen and sulfur resulting from atmospheric deposition. Thus, the Policy Assessment 

17 developed a new standard for aquatic acidification that is based on the use of chronic ANC as the 

18 ecological indicator as a component in the AAI.

19 The level of the standard would be defined in terms of a single, national value of the 

20 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.12 21 

22 The annual values of the AAI would be calculated based on the AAI equation using the assigned 

 

12 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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1 ecoregion-specific values for factors F1 through F4 and monitored annual average NOy and SOx 

2 concentrations. Since the AAI equation is based on chronic ANC as the ecological indicator, the 

3 level chosen for the standard would reflect a target chronic ANC value. As noted above, the 

4 assigned F factors for each ecoregion would be determined by EPA in the rulemaking to set the 

5 NAAQS, based on water quality and hydrology data, CMAQ modeling, the selected percentile 

6 that is used to identify a representative critical load within the ecoregion, and the selected level 

7 of the standard. The combination of the form of the standard, discussed above in section III.B, 

8 defined by the AAI equation and the assigned values of the F factors in the equation, other 

9 elements of the standard including the ambient air indicators (section III.A) and their averaging 

10 time (section III.C), and the level of the standard determines the allowable levels of NOy and 

11 SOx in the ambient air within each ecoregion. All of the elements of the standard together 

12 determine the degree of protection from adverse aquatic acidification effects associated with 

13 oxides of nitrogen and sulfur in the ambient air. The level of the standard plays a central role in 

14 determining the degree of protection provided and is discussed below.

15 The Policy Assessment focuses primarily on information that relates degrees of 

16 biological impairment associated with adverse ecological effects to aquatic ecosystems to 

17 alternative levels of ANC in reaching conclusions regarding the range of target ANC levels that 

18 is appropriate to consider for the level of the standard. The Policy Assessment develops the 

19 rationale for identifying a range of target ANC levels that is appropriate to consider by

20 addressing questions related to the following areas: (1) associations between ANC and pH levels 

21 to provide an initial bounding for the range of ANC values to be considered; (2) evidence that 

22 allows for the delineation of specific ANC ranges associated with varying degrees of severity of 

23 biological impairment ecological effects; (3) the role of ANC in affording protection against 
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1 episodic acidity; (4) implications of the time lag response of ANC to changes in deposition; (5) 

2 past and current examples of target ANC values applied in environmental management 

3 practices; and (6) data linking public welfare benefits and ANC.

4 1. Association Between pH Levels and Target ANC Levels 

5 As discussed above in section II.A and more fully in chapter 3 of the Policy Assessment, 

6 specific levels of ANC are associated with differing levels of risk of biological impairment in 

7 aquatic ecosystems, with higher levels of ANC resulting in lower risk of ecosystem impacts, and 

8 lower ANC levels resulting in risk of both higher intensity of impacts and a broader set of 

9 impacts. While ANC is not the causal agent determining biological effects in aquatic ecosystem,

10 it is a useful metric for determining the level at which a water body is protected against risks of 

11 acidification. There is a direct correlation between ANC and pH levels which, along with 

12 dissolved aluminum, are more closely linked to the biological causes of ecosystem response to 

13 acidification. 

14 Because there is a direct correlation between ANC and pH levels, the selection of target 

15 ANC levels is informed in part through information on effects of pH as well as direct studies of 

16 effects related to ANC. Levels of pH are closely associated with ANC in the pH range of 

17 approximately 4.5 to 7. Within this range, higher ANC levels are associated with higher pH 

18 levels. At a pH level of approximately 4.5, further reductions in ANC generally do not correlate 

19 with pH, as pH levels remain at approximately 4.5 while ANC values fall substantially. 

20 Similarly, at a pH value of approximately 7, ANC values can continue to increase with no 

21 corresponding increase in pH. As pH is the primary causal indicator of effects related to aquatic 

22 acidification, this suggests that ANC values below approximately -50 μeq/L (the apparent point 

23 in the relationship between pH and ANC where pH reaches a minimum) are not likely to result in 
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1 further damage. In addition, ANC values around and above approximately 100 μeq/L (the 

2 apparent region in the relationship where pH reaches a maximum) are not likely to confer 

3 additional protection. As a result, the initial focus in the Policy Assessment was on target ANC

4 values in the range of -50 to 100 μeq/L.

5 2. ANC Levels Related to Effects on Aquatic Ecosystems 

6 As discussed above in section II.A, the number of fish species present in a water body has 

7 been shown to be positively correlated with the ANC level in the water, with higher values 

8 supporting a greater richness and diversity of fish species. The diversity and distribution of 

9 phyto-zooplankton communities also are positively correlated with ANC. 

10 A summary of effects related to ANC ranges is shown above in Table II-1. Within the 

11 ANC range from approximately -50 to 100 μeq/L, linear and sigmoidal relationships are 

12 observed between ANC and ecosystem effects. On average, fish species richness is lower by one 

13 fish species for every 21 μeq/L decrease in ANC in Shenandoah National Park streams (ISA, 

14 section 3.2.3.4). As shown in Table II-1, ANC levels have been grouped into five categories 

15 related to expected ecological effects, including categories of acute concern ( <0 μeq/L), severe 

16 concern (0-20 μeq/L), elevated concern (20-50 μeq/L), moderate concern (50-100 μeq/L), and 

17 low concern (>100 μeq/L). This categorization is supported by a large body of research 

18 completed throughout the eastern U.S. (Sullivan et al., 2006).

19 Water bodies with ANC values less than or equal to 0 μeq/L at based flow are chronically 

20 acidic. Such ANC levels can lead to complete loss of species and major changes in the ability of 

21 water bodies to support diverse biota, especially in water bodies that are highly sensitive to 

22 episodic acidification. Based on the above considerations, the Policy Assessment has focused on 

23 target ANC levels no lower than 0 μeq/L.
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1 As discussed in the Policy Assessment, biota generally are not harmed when ANC values 

2 are >100 μeq/L, due to the low probability that pH levels will be below 7. In the Adirondacks,

3 the number of fish species also peaks at ANC values >100 μeq/L. This suggests that at ANC 

4 levels greater than 100 μeq/L, little risk from acidification exists in many aquatic ecosystems. At 

5 ANC levels below 100 μeq/L, overall health of aquatic communities can be maintained, although 

6 fish fitness and community diversity begin to decline. At ANC levels ranging from 100 down to 

7 50 μeq/L, there is increasing likelihood that the fitness of sensitive species (e.g., brook trout, 

8 zooplankton) will begin to decline. When ANC concentrations are below 50 μeq/L, the 

9 probability of acidification increases substantially, and negative effects on aquatic biota are 

10 observed, including large reductions in diversity of fish species and changes in the health of fish 

11 populations, affecting reproductive ability and fitness, especially in water bodies that are affected 

12 by episodic acidification. While there is evidence that ANC levels above 50 can confer 

13 additional protection from adverse ecological effects associated with aquatic acidification in 

14 some sensitive ecosystems, the expectation that such incremental protection from adverse effects 

15 will continue up to an ANC level of 100 is substantially reduced. The Policy Assessment 

16 concludes that the above considerations support a focus on target ANC levels up to a level 

17 greater than 50 μeq/L but below 100 μeq/L, such as up to a level of 75 μeq/L.

18 In considering the available scientific evidence, as summarized here and discussed in 

19 more detail in the ISA and REA, in its review of the second draft Policy Assessment, CASAC 

20 expressed the following views about the range of biological responses that corresponds to this 

21 range of ANC levels (i.e., 0-100 μeq/L):

22 There will likely be biological effects of acidification at higher ANC values 

23 within this range, and there are relatively insensitive organisms that are not 

24 impacted at ANC values at the low end of this range. Adverse effects of 

25 acidification on aquatic biota are fairly certain at the low end of this range of 
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1 ANC and incremental benefits of shifting waters to higher ANC become 

2 more uncertain at higher ANC levels. There is substantial confidence that 

3 there are adverse effects at ANC levels below 20 μeq/L, and reasonable 

4 confidence that there are adverse effects below 50 μeq/L. Levels of 50 

5 μeq/L and higher would provide additional protection, but the Panel has less 

6 confidence in the significance of the incremental benefits as the level 

7 increases above 50 μeq/L. (Russell and Samet, 2010b)

8

9 The Policy Assessment concludes that the above considerations, including the views of 

10 CASAC, provide support for focusing on target ANC levels in the range of 20 to 75 μeq/L.

11 3. Consideration of Episodic Acidity 

12 As discussed in the Policy Assessment, across the broad range of ANC values from 0 to 

13 100 μeq/L, ANC affords protection against the likelihood of decreased pH (and associated 

14 increases in Al) during long or short periods. In general, the higher the ANC within this range, 

15 the lower the probability of reaching low pH levels where direct effects such as increased fish 

16 mortality occur, as shown in Table 3-1 of the Policy Assessment. Accordingly, greater 

17 protection would be achieved by target chronic ANC values set high enough to avoid pH 

18 depression to levels associated with elevated risk. 

19 The specific relationship between ANC and the probability of reaching pH levels of 

20 elevated risk varies by water body and fish species. ANC levels below 20 μeq/L are generally 

21 associated with high probability of low pH, leading to death or loss of fitness of biota that are 

22 sensitive to acidification (US EPA, 2008, section 5.2.2.1; US EPA, 2009, section 5.2.1.2). At 

23 these levels, during episodes of high acidifying deposition, brook trout populations may

24 experience lethal effects. In addition, the diversity and distribution of zooplankton communities 

25 decline sharply at ANC levels below 20 μeq/L. Overall, there is little uncertainty that significant 

26 effects on aquatic biota are occurring at ANC levels below 20 μeq/L. 
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1 It is clear that at ANC levels approaching 0 μeq/L (Table II-1), there is significant 

2 impairment of sensitive aquatic ecosystems with almost complete loss of fish species. Avoiding 

3 ANC levels approaching 0 μeq/L is particularly relevant to episodic spikes in acidity that occur 

4 during periods of rapid snow melt and during and after major precipitation events. Since the 

5 ANC range considered in the Policy Assessment reflects average, long-term base flow values, it 

6 is appropriate to consider protecting against episodic drops in ANC values to a level as low as 0 

7 μeq/L. Staddard et al. (2003) noted on average a 30 μeq/L depression of ANC between spring 

8 and summer time values, indicating the need to maintain higher base flow ANC levels to protect 

9 against ANC levels below 0 μeq/L. The above considerations do not provide support for a target 

10 chronic ANC level as low as 0 μeq/L for a standard that would protect against significant harm to 

11 aquatic ecosystems, including harm from episodic acidification. The Policy Assessment 

12 concludes that these considerations also support a lower end of the range for consideration no 

13 lower than 20 μeq/L.

14 The CASAC agreed with this conclusion in its comments on the second draft Policy

15 Assessment (Russell and Samet, 2010b). CASAC noted that “there are clear and marked 

16 biological effects at ANC values near 0 μeq/L, so this is probably not an appropriate target 

17 value” for the AAI. With regard to the likelihood of impairment of aquatic ecosystems due to 

18 episodic acidification, in terms of specific target levels for chronic ANC, CASAC expressed the 

19 following view:

20 Based on surface waters studied in the Northeast, decreases in ANC 

21 associated with snowmelt [are] approximately 50 μeq/L. Thus, based on these 

22 studies, a long term ANC target level of 75 μeq/L would generally guard against 

23 effects from episodic acidification down to a level of about 25 μeq/L. (Russell and 

24 Samet, 2010b)

25

26

27
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1 4. Consideration of Ecosystem Response Time 

2 The Policy Assessment notes that when considering a standard level to protect against 

3 aquatic acidification, it is appropriate to take into account both the time period to recovery as 

4 well as the potential for recovery in acid-sensitive ecoregions. Ecosystems become adversely 

5 impacted by acidifying deposition over long periods of time and have variable time frames and 

6 abilities to recover from such perturbations. Modeling presented in the REA (US EPA, 2009,

7 section 4.2.4) shows the estimated ANC values for Adirondack lakes and Shenandoah streams 

8 under pre-acidification conditions and indicates that for a small percentage of lakes and streams, 

9 natural ANC levels would have been below 50 μeq/L. Therefore, for these water bodies, 

10 reductions in acidifying deposition are not likely to achieve an ANC of 50 μeq/L or greater. 

11 Conversely, for some lakes and streams the level of perturbation from long periods of acidifying 

12 deposition has resulted in very low ANC values compared to estimated natural conditions. For 

13 such water bodies, the time to recovery would be largely dependent on future inputs of acidifying 

14 deposition.

15 Setting a standard level in terms of a target chronic ANC level is based on the long-term 

16 response of aquatic ecosystems. The time required for a water body to achieve the target ANC 

17 level -- given a decrease in ambient air concentrations of NOy and SOx and related acidifying 

18 deposition such that the critical load for a target ANC is not exceeded – is often decades if not 

19 centuries. In recognition of the potential public welfare benefits of achieving the target ANC in 

20 a shorter time frame, the concept of target loads had been developed. Target loads represent the 

21 depositional loading that is expected to achieve a particular level of the ecological indicator by a 

22 given time. For example, to achieve an ANC level of 20 μeq/L by 2030, it might be necessary to 

23 specify a higher target ANC level of, for example, 50 μeq/L, such that the depositional loading 
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1 would be reduced more quickly than would occur if the depositional loading was based on 

2 achieving a target ANC level of 20 μeq/L as a long-term equilibrium level. In this example, the 

3 target ANC of 50 μeq/L would ultimately be realized many years later.

4 The above considerations have implications for selecting an appropriate standard level, in 

5 that the standard level affects not only the ultimate degree of protection that would be afforded 

6 by the standard, but also the time frame in which such protection would be realized. However, 

7 the Policy Assessment recognizes that there is a great deal of heterogeneity in response times 

8 among water bodies and that there is only very limited information from dynamic modeling that 

9 would help to quantify recovery time frames in areas across the country. As a consequence, 

10 quantification of a general relationship between critical loads associated with a specific long11 term target ANC level and target loads associated with achieving the target ANC level within a 

12 specific time frame is not currently possible. Thus, while the time frame for recovery is an 

13 important consideration in selecting an appropriate range of levels to consider, the Policy

14 Assessment concludes that it can only be considered in a qualitative sense at this time.

15 5. Prior Examples of Target ANC Levels 

16 A number of regional organizations, states, and international organizations have 

17 developed critical load frameworks to protect against acidification of sensitive aquatic 

18 ecosystems. In considering the appropriate range of target ANC levels for consideration in this 

19 review, it is informative to evaluate the target ANC levels selected by these different 

20 organizations, as well as the rationale provided in support of the selected levels. Chapter 4 of the 

21 Policy Assessment provides a detailed discussion of how critical loads have been developed and 

22 used in other contexts. Specific target values and their rationales are summarized below.
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1 The UNECE has developed critical loads in support of international emissions reduction 

2 agreements. As noted in chapter 4 of the Policy Assessment, critical loads were established to 

3 protect 95 percent of surface waters in Europe from an ANC less than 20 μeq/L based on 

4 protection of brown trout. Individual countries have set alternative ANC targets; for example, 

5 Norway targets an ANC of 30 μeq/L based on protection of Atlantic salmon.

6 Several states have established target ANC or pH values related to protection of lakes and 

7 streams from acidification. While recognizing that some lakes in the Adirondacks will have a 

8 naturally low pH, the state of New York has established a target pH value of 6.5 for lakes that 

9 are not naturally below 6.5. As noted above, this level is associated with an ANC value that is 

10 likely to be between 20 and 50 μeq/L or possibly higher. New Hampshire and Vermont have set 

11 ANC targets of 60 μeq/L and 50 μeq/L, respectively. Tennessee has established site-specific 

12 target ANC values based on assessments of natural acidity, with a default value of 50 μeq/L

13 when specific data are not available.

14 Taken together, these policy responses to concerns about ecological effects associated 

15 with aquatic acidification indicate that target ANC values between 20 and 60 μeq/L have been 

16 selected by states and other nations to provide protection of lakes and streams in some of the 

17 more sensitive aquatic ecosystems.

18 6. Consideration of Public Welfare Benefits

19 The point at which effects on public welfare become adverse is not defined in the CAA. 

20 Characterizing a known or anticipated adverse effect to public welfare is an important 

21 component of developing any secondary NAAQS. According to the CAA, welfare effects 

22 include:

23 ...effects on soils, water, crops, vegetation, manmade materials, animals, 

24 wildlife, weather, visibility, and climate, damage to and deterioration of 
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1 property, and hazards to transportation, as well as effect on economic values and 

2 on personal comfort and well-being, whether caused by transformation, 

3 conversion, or combination with other air pollutants. (CAA, section 302(h)).

4

5 Consideration of adversity to public welfare in the context of the secondary NAAQS for oxides 

6 of nitrogen and sulfur can be informed by information about losses in ecosystem services 

7 associated with acidifying deposition and the potential economic value of those losses, as

8 summarized above in section II.C and discussed more fully in chapter 4 of the Policy 

9 Assessment. 

10 Ecosystem service losses at alternative ANC levels are difficult to enumerate. However, 

11 in general there are categories of ecosystem services, discussed in chapter 4 of the Policy

12 Assessment, that are related to the specific ecosystem damages expected to occur at alternative 

13 ANC levels. Losses in fish populations due to very low ANC (below 20 μeq/L) are likely 

14 associated with significant losses in value for recreational and subsistence fishers. Many acid 

15 sensitive lakes are located in areas with high levels of recreational fishing activity. For example, 

16 in the northeastern U.S., where nearly 8 percent of lakes are considered acidic, more than 9 

17 percent of adults participate in freshwater fishing, with an estimated value of approximately $5 

18 billion in 2006. This suggests that improvements in lake fish populations may be associated with 

19 significant recreational fishing value.

20 As discussed in the Policy Assessment, inland surface waters also provide cultural 

21 services such as aesthetic and existence value and educational services. To the extent that 

22 piscivorous birds and other wildlife are harmed by the absence of fish in these waters, hunting 

23 and birdwatching activities are likely to be adversely affected. A case study of the value to New 

24 York residents of improving the health of lakes in the Adirondacks found significant willingness 

25 to pay for those improvements. When scaled to evaluate the improvement in lake health from 
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1 achieving ANC values of 20 to 50 μeq/L, the study implies benefits to the New York population 

2 roughly on the order of $600 million per year (in constant 2007$). The survey administered in 

3 this study recognized that participants were thinking about the full range of services provided by

4 the lakes in question – not just the recreational fishing services. Therefore the estimates of 

5 willingness to pay include resident’s benefits for potential hunting and birdwatching activities

6 and other ancillary services. These results are just for New York populations. The Policy

7 Assessment concludes that if similar benefits exist for improvements in other acid sensitive 

8 lakes, the economic value to U.S. populations could be very substantial, suggesting that, at least 

9 by one measure of impact on public welfare, impacts associated with ANC less than 50 μeq/L

10 may be adverse to public welfare.

11 7. Summary of Alternative Levels

12 Based on all the above considerations, the Policy Assessment concludes that 

13 consideration should be given to a range of standard levels from 20 to 75 μeq/L. The available 

14 evidence indicates that target ANC levels below 20 μeq/L would be inadequate to protect against 

15 substantial ecological effects and potential catastrophic loss of ecosystem function in some 

16 sensitive aquatic ecosystems. While ecological effects occur at ANC levels below 50 μeq/L in 

17 some sensitive ecosystems, the degree and nature of those effects are less significant than at 

18 levels below 20 μeq/L. Levels at and above 50 μeq/L would be expected to provide additional 

19 protection, although uncertainties regarding the potential for additional protection from adverse 

20 ecological effects are much larger for target ANC levels above about 75 μeq/L, as effects are 

21 generally appreciably less sensitive to changes in ANC at such higher levels.

22 In reaching this conclusion in the Policy Assessment, consideration was given to the 

23 extent to which a target ANC level within this range would protect against episodic as well as 
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1 long-term ecological effects. Levels in the mid- to upper part of this range would be expected to 

2 provide greater protection against short-term, episodic peaks in aquatic acidification, while lower 

3 levels within this range would give more weight to protection from long-term rather than 

4 episodic acidification. Similarly, levels in the mid- to upper part of this range would be expected 

5 to result in shorter time periods for recovery given the lag in ecosystem response in some 

6 sensitive ecosystems relative to levels in the lower part of this range. The Policy Assessment 

7 also note that this range encompasses target ANC values that have been established by various 

8 States and regional and international organizations to protect against acidification of aquatic 

9 ecosystems.

10 The Policy Assessment recognizes that the level of the standard together with the other 

11 elements of the standard, including the ambient air indicators, averaging time, and form, 

12 determine the overall protectiveness of the standard. Thus, consideration of a standard level 

13 should reflect the strengths and limitations of the evidence and assessments as well as the 

14 inherent uncertainties in the development of each of the elements of the standard. The 

15 implications of considering alternative standards, defined in terms of alternative combinations of 

16 levels and percentile values that are a critical component of factor F1 in the form of the standard, 

17 are discussed below in section III.E. Key uncertainties in the various components of the standard 

18 are summarized and considered below in section III.F. 

19 E. Combined Alternative Levels and Forms

20 To provide some perspective on the implications of various alternative multi-pollutant, 

21 AAI-based standards, the Policy Assessment presented the number of acid-sensitive ecoregions 

22 that would likely not meet various sets of alternative standards. The alternative standards 

23 considered were based on combinations of alternative target ANC levels, within the range of 20 
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1 to 75 μeq/L, and alternative forms, characterized by alternative representative percentiles within 

the range of the 70th to 90th 2 percentile. These alternative standards are also defined in terms of 

3 the other elements of the standard: ambient air indicators NOy and SOx, discussed above in 

4 section III.A; other elements of the form of the standard, including ecoregion-specific values for 

5 factors F1 through F4 in the AAI equation, discussed above in section III.B.5; and an annual 

6 averaging time for NOy and SOx, discussed above in section III.C. With regard to the averaging 

7 time, the assessment did not consider multi-year averaging of the calculated annual AAI values 

8 due to data limitations, including, for example, the lack of CMAQ modeling for multiple 

9 consecutive years. In this assessment, we characterize an ecoregion as likely not meeting a given 

10 alternative standard if the calculated AAI value is less than the target ANC level of the standard, 

11 recognizing that higher AAI values are more protective than lower values.

12 The results of this assessment are presented in Table 7-1a-d in the Policy Assessment for 

13 a subset of ecoregions including those characterized as acid sensitive. Calculated annual AAI 

14 values at the ecoregion level are shown for each alternative standard considered. Based on these 

15 AAI values, Table 7-2 in the Policy Assessment summarizes the number of acid-sensitive 

ecoregions that would likely not meet each of the alternative standards considered. 13 16 Calculated 

17 AAI values for all ecoregions categorized as relatively non-acid sensitive are shown in Table D-5 

18 in Appendix D of the Policy Assessment. In all cases, these relatively non-acid sensitive 

19 ecoregions were estimated to meet all of the alternative standards considered in this assessment.

20 As described above, the AAI values presented in Table 7-1a-d of the Policy Assessment 

21 are based in part on data from 2005 CMAQ model simulations, which was used to generate 

 

13 Tables 7-1a-d and 7-2 in the Policy Assessment present assessment results for 29 ecoregions that had been 

initially characterized as acid sensitive. Subsequently, based on a broader set of criteria used to characterize 

ecoregions as acid sensitive, as discussed above in section III.B.5.a, the set of ecoregions characterized as acid 

sensitive was narrowed to include 22 ecoregions.
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1 values for F2 through F4 in the AAI equation, as well as to estimate annual average ambient air 

2 concentrations of NOy and SOx that reflect recent air quality in the absence of currently 

3 available monitored concentrations in sensitive ecoregions across the country. Water quality and 

4 hydrology data from water bodies within each ecoregion were also used in calculating the AAI 

5 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 6 percentile critical load representative 

7 of the ecoregion used in calculating the F1 factor for the ecoregion. As expected, the number of 

8 ecoregions that likely would not meet alternative standards increases with increasing percentile 

9 values and target ANC levels (US EPA, 2011, Table 7-2). Out of 22 acid-sensitive ecoregions, 

10 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 11 percentile) to 4 for the least 

protective alternative standard (20 μeq/L, 70th 12 percentile). It is apparent that both the percentile 

13 and the level chosen have a strong influence, over the ranges considered, in determining the 

14 number of areas that would likely not meet this set of alternative standards. 

15 The Policy Assessment observes that there is one grouping of these acid-sensitive 

16 ecoregions that would likely not meet almost all combinations of level and form under 

17 consideration (US EPA, 2011, Table 7-2 and Appendix D). This group is made up of southern 

18 Appalachian mountain areas, including North Central Appalachians, 5.3.3; Ridge and Valley, 

19 8.4.1; Central Appalachians, 8.4.2; Blue Ridge, 8.4.4; and Southwestern Appalachians, 8.4.9. In 

20 addition, these ecoregions exhibit the highest amounts of exceedance relative to alternative 

21 standards.

22 The Northern Appalachian and Atlantic Maritime Highlands (5.3.1), which includes the 

23 Adirondacks, and the Northern Lakes and Forests (5.2.1) of the upper midwest exhibit similar 
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1 patterns with respect to in the role of level and percentile in identifying regions not likely to meet 

2 alternative standards, although there are considerably fewer cases compared to the regions in the 

3 Appalachians.

4 In the mountainous west, the Sierra Nevada (6.2.12), Idaho Batholith (6.2.15) and the 

5 Cascades (6.2.7) ecoregions likely would not meet alternative standards in fewer cases relative to 

6 eastern regions, with the Sierra Nevada ecoregion exhibiting relatively greater sensitivity

7 compared to all western regions. Only in the upper part of the ranges of level and percentile do 

8 regions in the northern and central Rockies likely not meet alternative standards.

9 In considering these findings, the Policy Assessment observes that the standard as 

10 defined by the AAI behaves in an intuitively logical manner. That is, an increase in ecoregions 

11 likely not to meet the standard is associated with higher alternative levels and percentiles, both of 

12 which contribute to a lower regionally representative critical load. Moreover, the areas of known 

13 adverse aquatic acidification effects are identified, mostly in high elevation regions or in the 

14 northern latitudes -- the Adirondacks, Shenandoahs, northern midwest lakes and the mountainous 

15 west. These results reflect the first application of a nationwide model that integrates water 

16 quality and atmospheric processes at a national scale and provides findings that are consistent 

17 with our basic understanding of the extent of aquatic acidification across the U.S. What is 

18 particularly noteworthy is that this model is not initialized with a starting ANC based on water 

19 quality data, which likely would result in a reproduction of water quality observations. Rather, 

20 this standard reflects the potential of the changes in atmospheric concentrations of NOy and SOx 

21 to induce long-term sustained changes in surface water systems. The Policy Assessment notes 

22 that the fact that the patterns of adversity based on applying this standard are commensurate with 
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1 what is observed in surface water systems provides confidence in the basic underlying 

2 formulation of the standard.

3 The Policy Assessment notes that the Appalachian mountain regions merit further 

4 inspection as they stand out as areas with the largest relative exceedances from a national 

5 perspective. Water quality data from these regions as well as an emissions sensitivity CMAQ 

6 simulation were considered to better understand the simulated behavior of these regions. The 

7 maps and tables in appendix D of the Policy Assessment include paired comparisons of the 

8 CMAQ 2005 and emissions sensitivity simulations. The emissions sensitivity simulation reflects 

9 domain-wide reductions in NOy and SOx emissions of 48% and 42%, respectively, relative to 

10 2005 base year emissions. The Policy Assessment assumes that this emissions sensitivity

11 simulation is indicative of future conditions. 

12 The emissions sensitivity results project that many of the regions that likely would not 

13 meet the alternative standards based on recent air quality, especially at alternative levels of 20 

14 and 35 μeq/L, would likely meet such standards in the future year scenario for the Appalachian

15 mountain regions. It is apparent that the AAI calculations are especially sensitive to changes in 

16 SOx emissions as the Appalachian regions have the highest SOx concentrations and deposition 

17 rates (US EPA, 2011,section 2), and the AAI equation responds as expected to modeled 

18 reductions in SOx. The emissions sensitivity scenario is a prospective application of the 

19 standard, in the sense that rules derived from the air quality management process result in 

20 reductions of NOy and SOx emissions. Expected emission changes over the next two decades 

21 should be far greater than the 42 and 48% SOx and NOy reductions used in this analysis, with a 

22 consequent further reduction in areas that would likely not meet alternative standards.
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1 The Appalachian mountain regions generally have low DOC levels, average runoff rates, 

2 moderately low base cation supply and highly elevated sulfate concentrations. Collectively, 

3 those attributes do not suggest naturally acidic conditions as the availability of anthropogenic 

4 contributions of mineral acids is likely responsible for observed low ANC values in those 

5 regions.

6 The Policy Assessment notes the Sierra Nevada region as an interesting case study, as it 

7 has some of the lowest critical load values nationally (US EPA, 2011, Table D-3). Water quality

8 data indicate extremely low sulfate, as expected given the relatively low SO2 emissions in the 

9 western U.S. Extremely low base cation supply and low Neco, which mitigate the effect of 

10 nitrogen deposition, explain the low critical load values. Low Neco values appear to associate 

11 well with high elevation western U.S. regions, perhaps reflecting the more arid and reduced 

12 vegetation density relative to eastern U.S. regions. The proximity to high level nitrogen 

13 emissions combined with very low base cation supply explains the cases where the Sierra region 

14 likely does not meet alternative standards. Because Neco values are low in the Sierras, the 

15 system responds effectively to reductions of NOx emissions, as illustrated in the maps and tables 

16 of Appendix D of the Policy Assessment. Although Neco affords protection from the acidifying 

17 effects of nitrogen deposition, the availability of excessive nitrogen neutralization capacity also 

18 means that reductions in nitrogen are not as effective as reductions in SOx in reducing the 

19 calculated AAI.

20 In reviewing these results, the Policy Assessment observes that the analysis of the 

21 alternative combinations of level and form presented provide context for considering the impact 

22 of different standards. Since the AAI equation has been newly developed in the Policy

23 Assessment, these examples of estimated exceedances help to address the question of whether 
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1 the AAI equation responds in a reasonable manner with regard to identifying areas of concern 

2 and to prospective changes in atmospheric conditions likely to result from future emissions 

3 reduction strategies. The Policy Assessment concludes that the behavior of the AAI calculations 

4 is both reasonable and explainable, which the Policy Assessment concludes serves to increase 

5 confidence in considering a standard defined in terms of the AAI.

6 F. Characterization of Uncertainties

7 This section summarizes discussions of the results of analyses and assessments, presented 

8 more fully in the Policy Assessment (US EPA, 2011, section 7.6 and Appendices F and G), 

9 intended to address the relative confidence associated with the linked atmospheric-ecological 

10 effects system described above. An overview of uncertainties is presented in the context of the 

11 major structural components underlying the standard, as well as with regard to areas of relatively 

12 high uncertainty. The section closes with a discussion of data gaps and uncertainties associated 

13 with the use of ecological and atmospheric modeling to specify the factors in the AAI equation, 

14 which can be used to guide future field programs and longer-term research efforts. 

15 1. Overview of Uncertainty

16 As discussed in the Policy Assessment (US EPA, 2011,Table 7-3), there is relatively low 

17 uncertainty with regard to the conceptual formulation of the overall structure of the AAI-based 

18 standard that incorporates the major associations linking biological effects to air concentrations. 

19 Based on the strength of the evidence that links species richness and mortality to water quality, 

20 the associations are strongly causal and without any obvious confounding influence. The strong 

21 association between the ecosystem indicator (ANC) and the causative water chemistry species 

22 (dissolved aluminum and hydrogen ion) reinforces the confidence in the linkage between 

23 deposition of nitrogen and sulfur and effects. This strong association between ANC and effects 
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1 is supported by a sound mechanistic foundation between deposition and ANC. The same 

2 mechanistic strength holds true for the relationship between ambient air levels of nitrogen and 

3 sulfur and deposition, which completes the linkage from ambient air indicators through 

4 deposition to ecological effects. 

5 There are relatively higher uncertainties, however, in considering specific elements

6 within the structure of an AAI-based standard, including the deposition of SOx, NOy, and NHx 

7 as well as the critical load-related component, each of which can vary within and across 

8 ecoregions. Overall system uncertainty relates not just to the uncertainty in each such element, 

9 but also to the combined uncertainties that result from linking these elements together within the 

10 AAI-based structure. Some of these elements – including, for example, dry deposition, pre11 industrial base cation production, and reduced nitrogen deposition – are estimated with less 

12 confidence than other elements (US EPA, 2011, Table 7.3). The uncertainties associated with all 

13 of these elements, and the combination of these elements through the AAI equation, are 

14 discussed below and in the following sections related to measured data gaps and modeled 

15 processes for both air quality and water quality.

16 The lack of observed dry deposition data is constrained by resources and the lack of 

17 efficient measurement technologies. Progress in reducing uncertainties in dry deposition will 

18 depend on improved atmospheric concentration data and direct deposition flux measurements of 

19 the relevant suite of NOy and SOx species. 

20 Pre-industrial base cation productivity by definition is not observable. Contemporary 

21 observations and inter-model comparisons are useful tools that would help reduce the uncertainty

22 in estimates of preindustrial base cation productivity used in the AAI equation. In characterizing 

23 contemporary base cation flux using basic water quality measurements (i.e., major anioni and 
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1 cation species as defined in equation 2.11 in the Policy Assessment), it is reasonable to assume 

2 that a major component of contemporary base cation flux is associated with pre-industrial 

3 weathering rates. To the extent that multiple models converge on similar solutions, greater 

4 confidence in estimating pre-industrial base cation production would be achieved.

5 Characterization of reduced nitrogen (NHx) deposition has been evolving over the last 

6 decade. The relatively high uncertainty in characterizing NHx deposition is due to both the lack 

7 of field measurements and the inherent complexity of characterizing NHx with respect to source 

8 emissions and dry deposition. Because ammonia emissions are generated through a combination 

9 of man-made and biological activities, and ammonia is semi-volatile, the ability to characterize 

10 spatial and temporal distributions of NHx concentrations and deposition patterns is challenging. 

11 While direct measurement of NHx deposition is resource intensive because of the diffuse nature 

12 of sources (i.e., area-wide and not point sources), there have been more frequent deposition flux 

13 studies, relative to other nitrogen species, that enable the estimation of both emissions and dry

14 deposition. Also, while ammonia has a relatively high deposition velocity and traditionally was 

15 thought to deposit close to the emissions release areas, the semi-volatile nature of ammonia 

16 results in re-entrainment back into the lower boundary layer resulting in a more dispersed 

17 concentration pattern exhibiting transport type characteristics similar to longer lived atmospheric 

18 species. These inherent complexities in source characterization and ambient concentration 

19 patterns raise the uncertainty level of NHx in general. However, the Policy Assessment notes 

20 that progress is being made in measuring ammonia with cost efficient samplers and anticipates 

21 the gradual evolution of a spatially robust ammonia sampling network that would help support 

22 analyses to reduce underlying uncertainties in NHx deposition. Also, from an aquatic 

23 acidification perspective, NHx is not as important a driver as NOy and SOx in the mountainous 
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1 areas in the eastern U.S. However, the relative importance of NHx is likely to increase over 

2 time, in light of air quality rules in place designed to reduce emissions of NOy and SOx.

3 2. Uncertainties Associated with Data Gaps

4 In summarizing uncertainties with respect to available measurement data and the use of 

5 ecological and atmospheric models, the Policy Assessment indentified data gaps and model 

6 uncertainties in relative terms by comparing, for example, the relative richness of data between 

7 geographic areas or environmental media. With regard to relevant air quality measurements, the 

8 Policy Assessment notes that such measurements are relatively sparse in the western U.S. While 

9 the spatial extent of CASTNET coverage has gradually incorporated western U.S. locations with 

10 support from the National Parks Service (NPS), the relative density of monitoring sites is much 

11 less than that in the eastern U.S. This relative disparity in spatial density of monitors is 

12 exacerbated as air quality patterns in the mountainous west generally exhibit greater spatial 

13 heterogeneity due to dramatic elevation gradients that impact meteorology and air mass flow 

14 patterns. Similarly, water quality data coverage is far more comprehensive in the eastern U.S. 

15 relative to the west

16 Measurements of NOy notably are lacking in both eastern and western acid-sensitive 

17 ecoregions. This adds uncertainty to the use of the AAI equation as the lack of NOy data limits 

18 efforts to evaluate air quality modeling of NOy that is the basis for quantifying factor F3 in the 

19 AAI equation. The lack of NOy measurements also limits efforts to characterize the variability

20 and representativeness of modeled NOy concentrations within and across ecoregions. Currently, 

21 the Agency’s ability to define the protection likely to be afforded by alternative standards (in 

22 terms of alternative levels and percentiles) is compromised by the lack of a full set of ambient air 
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1 quality indicator measurements, notably including NOy, throughout sensitive ecoregions across 

2 the U.S. 

3 Further, obtaining measurement of the dominant species that comprise NOy (HNO3, true 

4 NO2, NO, p–NO3, and PAN) would be useful to evaluate performance of NOy samplers. 

5 Beyond the more well known dominant components of NOy, research efforts would be needed to 

6 characterize total reactive nitrogen that may include significant amounts of organically-bound 

7 nitrogen (beyond PAN) which is poorly understood with regard to emission sources and 

8 concentration levels.

9 Field measurements of NHx have been extremely limited, but have begun to be enhanced 

10 through the National Atmospheric Deposition Program’s (NADP) passive ammonia network 

11 (AMoN). AMoN measures ammonia at over 50 sites, with more than 35 at CASTNET locations. 

12 Enhanced spatial coverage of reduced nitrogen measurements, particularly to understand within 

13 and across ecoregion variability, and the inclusion of some continuous observations would 

14 provide a better understanding of the uncertainty in the F2 factor in the AAI equation and of the

15 representativeness of modeled NHx deposition within and across ecoregions.

16 With regard to water quality data, the Policy Assessment notes that such data are 

17 typically limited relative to air quality data sets, and are also relatively sparse in the western U.S. 

18 The TIME/LTM water quality sampling program in the eastern U.S. (as described in chapter 2 

19 of the Policy Assessment) is an appropriate complement to national air monitoring programs as it 

20 affords consistency across water bodies in terms of sampling frequency and analysis protocols. 

21 Consideration should be given to extending the TIME/LTM design to all acid sensitive 

22 ecoregions, with priority for areas in the western mountains that are data limited and showing 

23 initial signs of adversity particularly with respect to aquatic acidification. The lack of a 
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1 regulatory requirement for TIME/LTM often jeopardizes funding support of this resource that is 

2 especially valuable and cost effective. While there are several state and local agency water 

3 quality data bases, it is unclear the extent to which differences in sampling, chemical analysis 

4 and reporting protocols would impact the use of such data for the purpose of better 

5 understanding the degree of protectiveness that would be afforded by an AAI-based standard 

6 within sensitive ecoregions across the country. In addition, our understanding of water quality in 

7 Alaska and Hawaii and the acid sensitivity of their ecoregions is particularly limited. 

8 Water quality data and modeling support the standard setting process. As more water 

9 bodies are sampled, the critical load data bases would expand, enabling clearer delineation of 

ecoregion representative critical loads in terms of the nth 10 percentile. This would provide more 

11 refined characterization of the degree of protection afforded by a given standard. Longer term, 

12 the availability of water quality trend data (annual to monthly sampled) would support 

13 accountability assessments that examine if an ecoregion’s response to air management efforts is 

14 as predicted by earlier model forecasting. The most obvious example is the long-term response 

15 of water quality ANC change to changes in calculated AAI, deposition, ambient NOy and SOx 

16 concentrations, and emissions. In addition, water quality trends data provide a basis for 

17 evaluating and improving the parameterizations of processes in critical load models applied at 

18 the ecoregion scale related to nitrogen retention and base cation supply. A better understanding

19 of soil processes, especially in the southern Appalachians, would enhance efforts to examine the 

20 variability within ecoregions of the soil-based adsorption and exchange processes which 

21 moderate the supply of major cations and anions to surface waters and strongly influence the 

22 response of surface water ANC to changes in deposition of nitrogen and sulfur.
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1 3. Uncertainties in Modeled Processes

2 As discussed in the Policy Assessment, from an uncertainty perspective, gaps in field 

3 measurement data are related to uncertainties in modeled processes and in the specific 

4 application of such models. As noted above, processes that are embodied in an AAI-based 

5 standard are modeled using the CMAQ atmospheric model and steady state ecological models. 

6 These models are characterized in the Integrated Science Assessment as being well established 

7 and they have undergone extensive peer review. Nonetheless, the application of these models for 

8 purposes of specifying the factors in the AAI equation, on an ecoregion scale, is a new 

9 application that introduces uncertainties, as noted below, especially in areas with limited 

10 observational data that can be used to evaluate this specific application. Understanding 

11 uncertainties in relevant modeled process thus involves consideration of the uncertainties 

12 associated with applying each model as well as the combination of these uncertainties as the

13 models are applied in combination within the AAI framework.

14 With regard to the application of CMAQ for purposes of use in an AAI-based standard, 

15 the modeling of dry deposition has been identified as having a relatively high degree of 

16 uncertainty. Due to a combination of system complexity and resource constraints, there is no 

17 routine observational basis for directly comparing modeled dry deposition and measurements. 

18 Periodic dry deposition flux experiments covering a variety of vegetation, surfaces and 

19 meteorology across seasons would enable a more robust evaluation of modeled deposition of 

20 nitrogen and sulfur. Given the difficulty in acquiring dry deposition observations, it becomes 

21 especially important to evaluate the model’s ability to capture temporal and spatial ambient air 

22 patterns of individual nitrogen and sulfur species which are used to drive dry deposition 

23 calculations in models. For example, reducing a generally acknowledged positive bias in model-
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1 predicted SO2 relative to observations is especially relevant to the AAI-based standard, as SO2

2 deposition is a dominant contributor to total acidifying deposition in the eastern U.S. With 

3 respect to oxidized nitrogen, observations of individual NOy species are important as air quality 

4 models calculate the individual deposition of each species. The modeled transference ratios, 

5 TNOy and TSOx used in factors F3 and F4 rely on CMAQ’s ability to characterize both deposition 

6 and concentration. Consequently, a better understanding of the variability of these factors 

7 within and across ecoregions could be achieved by improved availability of measured ambient 

8 concentrations and deposition observations.

9 Steady state biogeochemical ecosystem modeling is used to develop critical load 

10 estimates that are incorporated in the AAI equation through factor F1. Consequently, the Policy

11 Assessment notes that an estimate of the temporal response of surface water ANC to deposition 

12 and air concentration changes is not directly available. Lacking a predicted temporal response 

13 impairs the ability to conduct accountability assessments down to the effects level. 

14 Accountability assessments would examine the response of each step in the emissions source 

15 through air concentration – deposition -- surface water quality – biota continuum. The steady

16 state assumption at the ecosystem level does not impair accountability assessments through the 

17 air concentration/deposition range of that continuum. However, in using steady state ecosystem 

18 modeling, several assumptions are made relative to the long-term importance of processes 

19 related to soil adsorption of major ions and ecosystem nitrogen dynamics. Because these models 

20 often were developed and applied in glaciated areas with relatively thin and organically rich 

21 soils, their applicability is relatively more uncertain in areas such as those in the non-glaciated 

22 clay-based soil regions of the central Appalachians. Consequently, it is desirable to develop the 

23 information bases to drive simple dynamic ecosystem models that incorporate more detailed 
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1 treatment of subsurface processes, such as adsorption and exchange processes and sulfate 

2 absorption.

3 4. Applying Knowledge of Uncertainties

4 An understanding of the relative uncertainties in a system assists in setting priorities for 

5 data collection efforts and research, with the expectation that such efforts would reduce 

6 uncertainties over time and afford greater confidence in applications of an AAI-based standard. 

7 Because of the uniquely wide breadth of pollutants and environmental media addressed by an

8 AAI-based multi-pollutant standard, there are a wide range of uncertainties that are important to 

9 consider relative to single pollutant standards that typically address only direct effects of ambient 

10 air exposures. For an AAI-based standard, a reduction of the uncertainties across the various 

11 modeled processes at the ecoregion scale would lead to greater confidence in the degree of 

12 protection afforded by the standard. 

13 The Policy Assessment notes that there is generally low uncertainty with regard to the 

14 conceptual development and related major components of this standard. In recognizing the 

15 scientific soundness of the basic structure of this standard, the Policy Assessment notes that 

16 future efforts would be appropriately directed at expanding the availability of relevant data for 

17 ecoregion-specific evaluation and application of the relevant modeling of ecological and 

18 atmospheric processes, as identified above. Such efforts would further support consideration of 

19 an AAI-based standard and would guide field studies and analyses designed to improve the 

20 longer-term confidence in such a standard.

21 G. CASAC Advice

22 The CASAC has advised EPA concerning the Integrated Science Assessment, the Risk 

23 and Exposure Assessment, and the Policy Assessment. CASAC has endorsed EPA’s 
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1 interpretation of the science embodied in the Integrated Science Assessment and the assessment 

2 approaches and conclusions incorporated in the Risk and Exposure Assessment.

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

4 providing its recommendations on the review of the new multi-pollutant standard developed in 

5 that document and discussed above (Russell and Samet, 2011a). In so doing, CASAC has 

6 expressed general support for the conceptual framework of the standard based on the underlying 

7 scientific information, as well as for the conclusions in the Policy Assessment with regard to 

8 indicators, form, averaging time, and level of the standard that are appropriate for consideration 

9 by the Agency in reaching decisions on the review of the secondary NAAQS for oxides of 

10 nitrogen and sulfur:

11 The final Policy Assessment clearly sets out the basis for the recommended ranges 

12 for each of the four elements (indicator, averaging time, level and form) of a 

13 potential NAAQS that uses ambient air indicators to address the combined effects 

14 of oxides of nitrogen and oxides of sulfur on aquatic ecosystems, primarily 

15 streams and lakes. As requested in our previous letters, the Policy Assessment also 

16 describes the implications of choosing specific combinations of elements and 

17 provides numerous maps and tabular estimates of the spatial extent and degree of 

18 severity of NAAQS exceedances expected to result from possible combinations of 

19 the elements of the standard.

20

21 We believe this final PA is appropriate for use in determining a secondary

22 standard to help protect aquatic ecosystems from acidifying deposition of oxides 

23 of sulfur and nitrogen. EPA staff has done a commendable job developing the 

24 innovative Aquatic Acidification Index (AAI), which provides a framework for a 

25 national standard based on ambient concentrations that also takes into account 

26 regional differences in sensitivities of ecosystems across the country to effects of 

27 acidifying deposition. (Russell and Samet, 2011a)

28

29 CASAC also recommended that as EPA moves forward in the regulatory process “some 

30 attention should be given to our residual concern that the available data may reflect the more 

31 sensitive water bodies and thus, the selection of percentiles of waterbodies to be protected could 
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1 be conservatively biased” (Russell and Samet, 2011a). In additional, CASAC found some 

2 improvements could be made to the uncertainty analysis, as noted below.

3 With respect to indicators, CASAC supports the use of SOx and NOy as ambient air indicators 

4 (discussed above in section III.A) and ANC as the ecological indicator (discussed above in 

5 section III.B.1):

6 The use of total reactive oxidized nitrogen (NOy) and sulfur oxides (SOx) as atmospheric 

7 indicators of oxidized nitrogen (N) and sulfur (S) atmospheric concentrations is well justified. 

8 The use in the Aquatic Acidification Index (AAI) of NOy and SOx as atmospheric indicators of 

9 N and S concentrations is useful and corresponds with other efforts by EPA. As we have stated 

10 previously, CASAC also agrees that acid neutralizing capacity (ANC) is the most appropriate 

11 ecological indicator of aquatic ecosystem response and resiliency to acidification. (Russell and 

12 Samet, 2011a)

13 With respect to the form of the standard (discussed above in section III.B), CASAC 

14 stated the following:

15 EPA has developed the AAI, an innovative “form” of the NAAQS itself that 

16 incorporates the multi-pollutant, multi-media, environmentally modified, 

17 geographically variable nature of SOx/NOy deposition-related aquatic 

18 acidification effects. With the caveats noted below, CASAC believes that this 

19 form of the NAAQS as described in the final Policy Assessment is consistent with 

20 and directly reflective of current scientific understanding of effects of acidifying 

21 deposition on aquatic ecosystems. (Russell and Samet, 2011a)

22

23 CASAC agrees that the spatial components of the form in the Policy Assessment 

24 are reasonable and that use of Omernick’s ecoregions (Level III) is appropriate for 

25 a secondary NAAQs intended to protect the aquatic environment from 

26 acidification . . . (Russell and Samet, 2011a)

27

28 The “caveats” noted by CASAC include a recognition of the importance of continuing to 

29 evaluate the performance of the CMAQ and ecological models to account for model 
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1 uncertainties and to make the model-dependent factors in the AAI more transparent. In addition, 

2 CASAC noted that the role of DOC and its effects on ANC would benefit from further 

3 refinement and clarification (Russell and Samet, 2011a). While CASAC expressed the view that 

4 the “division of ecoregions into ‘sensitive’ and ‘non-sensitive’ subsets, with a more protective 

5 percentile applied to the sensitive areas, also seems reasonable” (Russell and Samet, 2011a),

6 CASAC also noted that there was the need for greater clarity in specifying how appropriate 

7 screening criteria would be applied in assigning ecoregions to these categories. Further, CASAC 

8 identified potential biases in critical load calculations and in the regional representativeness of 

9 available water chemistry data, leading to the observation that a given percentile of the 

10 distribution of estimated critical loads may be protective of a higher percentage of surface waters 

11 in some regions (Russell and Samet, 2011a).

12 With respect to averaging time (discussed above in section III.C), CASAC stated the 

13 following:

14 Considering the cumulative nature of the long-term adverse ecological effects and 

15 the year-to-year variability of atmospheric conditions (mainly in the amount of 

16 precipitation), CASAC concurs with EPA that an averaging time of three to five 

17 years for the AAI parameters is appropriate. A longer averaging time would mask 

18 possible trends of AAI, while a shorter averaging time would make the AAI being 

19 more influenced by the conditions of the particular years selected. (Russell and 

20 Samet, 2011a)

21

22 With respect to level as well as the combination of level and form as they are presented 

23 as alternative standards (discussed above in sections III.D-E), CASAC stated the following:

24 CASAC agrees with EPA staff’s recommendation that the “level” of the 

25 alternative AAI standards should be within the range of 20 and 75 μeq/L. We also 

26 recognize that both the “level” and the form of any AAI standard are so closely

27 linked in their effectiveness that these two elements should be considered 

28 together. (Russell and Samet, 2011a)

29
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1 When considered in isolation, it is difficult to evaluate the logic or implications 

2 of selecting from percentiles (70th to 90th) of the distribution of estimated critical 

3 loads for lakes in sensitive ecoregions to determine an acceptable amount of 

4 deposition for a given ecoregion. However, when these percentile ranges are 

5 combined with alternative levels within the staff-recommended ANC range of 20 

6 to 75 microequivalents per liter (μeq/L), the results using the AAI point to the 

7 ecoregions across the country that would be expected to require additional 

8 protection from acidifying deposition. Reasonable choices were made in 

9 developing the form. The number of acid sensitive regions not likely to meet the 

10 standard will be affected both by choice of ANC level and the percentile of the 

11 distribution of critical loads for lakes to meet alternative ANC levels in each 

12 region. These combined recommendations provide the Administrator with a 

13 broad but reasonable range of minimally to substantially protective options for the 

14 standard. (Russell and Samet, 2011a)

15

16 CASAC also commented on EPA’s uncertainty analysis, and provided advice on areas 

17 requiring further clarification in the proposed rule and future research. CASAC found it 

18 “difficult to judge the adequacy of the uncertainty analysis performed by EPA because of lack of 

19 details on data inputs and the methodology used, and lack of clarity in presentation” (Russell and 

20 Samet, 2011a). In particular, CASAC identified the need for more thorough model evaluations 

21 of critical load and atmospheric modeling, recognizing the important role of models as they are 

22 incorporated in the form of the standard. In light of the innovative nature of the standard 

23 developed in the Policy Assessment, CASAC identified “a number of areas that should be the 

24 focus of further research” (Russell and Samet, 2011a). While CASAC recognized that EPA staff 

25 was able to address some of the issues in the Policy Assessment, they also noted areas “that 

26 would benefit from further study or consideration in potential revisions or modifications to the 

27 form of the standard.” Such research areas include “sulfur retention and mobilization in the 

28 soils, aluminum availability, soil versus water acidification and ecosystem recovery times.” 

29 Further, CASAC encouraged future efforts to monitor individual ambient nitrogen species, 
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1 which would help inform further CMAQ evaluations and the specification of model-derived 

2 elements in the AAI equation (Russell and Samet, 2011a).

3 H. Administrator’s Proposed Conclusions

4 Having concluded that the existing NO2 and SO2 secondary standards are neither 

5 sufficiently protective nor appropriate to address deposition-related effects associated with 

6 oxides of nitrogen and sulfur (section II.D above), the Administrator has considered whether it is 

7 appropriate at this time to set a new multi-pollutant standard for that purpose, with a structure 

8 that would better reflect the available science regarding acid deposition. In particular, she has 

9 focused on the new standard developed in the Policy Assessment and reviewed by CASAC, as 

10 discussed above. In so doing, the Administrator first considered the extent to which there is a 

11 scientific basis for development of a standard that would be requisite to protect public welfare, as 

12 provided in section 109(b) of the CAA, specifically with respect to deposition-related aquatic 

13 acidification in sensitive aquatic ecosystems, that would be neither more nor less stringent than 

14 necessary for this purpose. As discussed above, the Administrator notes that the Integrated 

15 Science Assessment concludes that the available scientific evidence is sufficient to infer a causal 

16 relationship between acidifying deposition of nitrogen and sulfur in aquatic ecosystems, and that 

17 the deposition of oxides of nitrogen and sulfur both cause such acidification under current 

18 conditions in the U.S. Further, the Integrated Science Assessment concludes that there are well19 established water quality and biological indicators of aquatic acidification as well as well20 established models that address deposition, water quality, and effects on ecosystem biota, and 

21 that ecosystem sensitivity to acidification varies across the country according to present and 

22 historic nitrogen and sulfur deposition as well as geologic, soil, vegetative, and hydrologic 

23 factors. Based on these considerations, the Administrator agrees with the conclusion in the 
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1 Policy Assessment, and supported by CASAC, that there is a strong scientific basis for 

2 development of a standard with the general structure presented in the Policy Assessment.

3 The Administrator also recognizes that the conceptual framework for an ecologically

4 relevant, multi-pollutant standard, which was initially explored in the Risk and Exposure 

5 Assessment and further developed in the Policy Assessment, builds on the information in the 

6 Integrated Science Assessment. She notes that the structure of the standard addresses the 

7 combined effects of deposition from oxides of nitrogen and sulfur by characterizing the linkages 

8 between ambient concentrations, deposition, and aquatic acidification, and that the structure of 

9 the standard takes into account relevant variations in these linkages across the country. She 

10 recognizes that while the standard is innovative and unique, the structure of the standard is well 

11 grounded in the science underlying the relationships between ambient concentrations of oxides 

12 of nitrogen and sulfur and the aquatic acidification related to deposition of nitrogen and sulfur 

13 associated with such ambient concentrations.

14 While the Administrator recognizes the strong scientific foundation for the structure of an

15 AAI-based standard, she also recognizes that the standard depends on atmospheric and 

16 ecological modeling, based on appropriate data, to specify the terms of an equation that 

17 incorporates the linkages between ambient concentrations, deposition, and aquatic acidification. 

18 This equation, which defines an aquatic acidification index (AAI), has the effect of translating 

19 spatially variable ecological effects into a potential national standard. With respect to 

20 establishing the specific terms of this equation, there are a number of inherent uncertainties and 

21 complexities that are relevant to the question of whether it is feasible to set a specific AAI-based

22 standard at this time that is requisite to protect public welfare without being either more or less 

23 stringent than necessary for this purpose. As discussed above, these uncertainties and 
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1 complexities generally relate not to the structure of the standard, but to the quantification of the 

2 various elements of the standard, such as the F factors discussed earlier in this section and their 

3 representativeness at an ecoregion scale. These uncertainties and complexities currently limit 

4 efforts to determine an appropriate form an level for such a standard, within the ranges of levels 

5 and forms identified in the Policy Assessment, and the representativeness of F factors in the AAI 

6 equation described above and in the Policy Assessment. These important uncertainties have 

7 been generally categorized as limitations in available field data as well as uncertainties that are 

8 related to reliance on the application of ecological and atmospheric modeling at the ecoregion 

9 scale to specify the various elements of the AAI.

10 With regard to data limitations, the Administrator observes that there are several 

11 important limitations in the available data upon which elements of the AAI are based. For 

12 example, while ambient measurements of NOy are made as part of a national monitoring 

13 network, the monitors are not located in locations that are representative of sensitive aquatic 

14 ecosystems. While air and water quality data are generally available in areas in the eastern U.S., 

15 there is relatively sparse coverage in mountainous western areas where a number of sensitive 

16 aquatic ecosystems are located. Further, even in areas where relevant data are available, small 

17 sample sizes impede efforts to characterize the representativeness of the available data, which 

18 was noted by CASAC as being of particular concern. Also, measurements of reduced forms of 

19 nitrogen are available from only a small number of monitoring sites, and emission inventories for 

20 reduced forms of nitrogen used in atmospheric modeling are subject to considerable uncertainty.

21 With regard to uncertainties related to the use of ecological and atmospheric modeling, 

22 the Administrator notes in particular that model results are difficult to evaluate due to a lack of 

23 relevant observational data. For example, relatively large uncertainties are introduced by a lack 
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1 of data with regard to pre-industrial environmental conditions and other parameters that are 

2 necessary inputs to critical load models that are the basis for factor F1 in the AAI equation. 

3 Also, observational data are not generally available to evaluate the modeled relationships 

4 between nitrogen and sulfur in the ambient air and associated deposition, which are the basis for 

5 the other factors (i.e., F2, F3, and F4) in the AAI equation.

6 In combination, these limitations and uncertainties result in a considerable degree of 

7 uncertainty as to how well the quantified elements of the AAI standard would predict the actual 

8 relationship between varying ambient concentrations of oxides of nitrogen and sulfur and steady 

9 state ANC levels across the distribution of water bodies within the various ecoregions in the U.S. 

10 Because of this, there is considerable uncertainty as to the actual degree of protectiveness that 

11 such a standard would provide, especially for acid-sensitive ecoregions. The Administrator 

12 recognizes that the AAI equation, with factors quantified in the ranges discussed above and 

13 described more fully in the Policy Assessment, generally performs well in identifying areas of 

14 the country that are sensitive to such acidifying deposition and indicates, as expected, that lower 

15 ambient levels of oxides of nitrogen and sulfur would lead to higher calculated AAI values. 

16 However, the uncertainties discussed here are critical for determining the actual degree of 

17 protection that would be afforded such areas by any specific target ANC level and percentile of 

18 water bodies that would be chosen in setting a new AAI-based standard, and thus for determining 

19 an appropriate AAI-based standard that meets the requirements of the CAA.

20 In considering these uncertainties, the Administrator notes that CASAC acknowledged 

21 that important uncertainties remain that would benefit from further study and data collection 

22 efforts, which might lead to potential revisions or modifications to the form of the standard 

23 developed in the Policy Assessment. She also notes that CASAC encouraged the Agency to 
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1 engage in future monitoring and model evaluation efforts to help inform the specification of 

2 model-derived elements in the AAI equation.

3 Based on the above considerations, the Administrator has determined that it is not 

4 appropriate to set a new multi-pollutant standard to address deposition-related effects of oxides 

5 of nitrogen and sulfur on aquatic acidification at this time. Setting a NAAQS generally involves 

6 consideration of the degree of uncertainties in the science and other information, such as gaps in 

7 the relevant data and, in this case, limitations in the evaluation of the application of relevant 

8 ecological and atmospheric models at an ecoregion scale. The issue here is identifying a specific 

9 form and level for the standard that would provide an appropriate level of public welfare 

10 protection, consistent with the requirements of the CAA, in light of the specific uncertainties and 

11 limitations that are discussed in detail throughout this preamble. As noted above, it is not a 

12 question of uncertainties about the scientific soundness of the structure of the AAI, but instead 

13 uncertainties in the quantification and representativeness of the elements of the AAI as they vary

14 in ecoregions across the country. At present, these uncertainties prevent an understanding of the 

15 degree of protectiveness that would be afforded to various ecoregions across the country by a 

16 new standard defined in terms of a specific nationwide target ANC level and a specific percentile 

17 of water bodies for acid-sensitive ecoregions, and thus prevent identification of an appropriate 

18 form and level for the standard. The Administrator has considered whether these uncertainties 

19 could be appropriately accounted for by choosing either a more or less protective target ANC 

20 level and percentile of water bodies than would otherwise be chosen if the uncertainties did not 

21 substantially limit the confidence that can appropriately be ascribed to the quantification of the 

22 AAI elements. However, in the Administrator’s judgment, the uncertainties are of such nature 

23 and magnitude that there is no reasoned way to choose such a specific nationwide target ANC 
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1 level or percentile of water bodies that would appropriately account for the uncertainties, since 

2 neither the direction nor the magnitude of change from the target level and percentile that would 

3 otherwise be chosen can reasonably be ascertained at this time. 

4 Based on the above considerations, the Administrator judges that the current limitations 

5 in relevant data and the uncertainties associated with specifying the elements of the AAI based 

6 on modeled factors are of such nature and degree as to prevent her from reaching a reasoned 

7 decision such that she is adequately confident as to what level and form (in terms of a selected 

8 percentile) of such a standard would provide any particular intended degree of protection of 

9 public welfare that the Administrator determined satisfied the requirements of the CAA. While 

10 acknowledging that CASAC supported moving forward to establish the standard developed in 

11 the Policy Assessment, the Administrator also observes that CASAC supported conducting 

12 further field studies that would better inform the continued development or modification of such 

13 a standard. Given the large uncertainties and complexities inherent in quantifying the elements 

14 of such a standard, largely deriving from the unprecedented nature of the standard under 

15 consideration in this review, and having fully considered CASAC’s advice, the Administrator 

16 provisionally concludes that it is premature to set a new, multi-pollutant secondary standard for 

17 oxides of nitrogen and sulfur at this time, and as such she is proposing not to set such a new 

18 secondary standard.

19 While it is premature to set such a multi-pollutant standard at this time, the Administrator 

20 determines that the Agency should undertake a field pilot program to gather additional data, and 

21 that it is appropriate that such a program be undertaken before, rather than after, reaching a 

22 decision to set such a standard. As described below in section IV, the purpose of the program is

23 to collect and analyze data so as to enhance our understanding of the degree of protectiveness 
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1 that would likely be afforded by a standard based on the AAI as developed in the Policy

2 Assessment. This will provide a better basis for the Administrator to identify a form and level of 

3 the standard that is requisite to protect the public welfare consistent with the CAA, specifically

4 with respect to the acidifying effects of deposition of oxides of nitrogen and sulfur. Data 

5 generated by this field program will also support development of an appropriate monitoring 

6 network that would work in concert with such a standard to result in the intended degree of 

7 protection. The data and analyses generated as a result of this program will serve to inform the 

8 next review of the NAAQS for oxides of nitrogen and sulfur. The information generated during 

9 the field program can also be used to help state agencies and EPA better understand how an 

10 AAI-based standard would work in terms of the implementation of such a standard.

11 Based on the above considerations, the Administrator is proposing not to set a new multi12 pollutant AAI-based secondary standard for oxides of nitrogen and sulfur in this review. In 

13 reaching this decision, the Administrator recognizes that the new NO2 and SO2 primary 1-hour 

14 standards set in 2010, while not ecologically relevant for a secondary standard, will nonetheless 

15 result in reductions in oxides of nitrogen and sulfur that will directionally benefit the 

16 environment by reducing NOy and SOx deposition to sensitive ecosystems. EPA is proposing to 

17 revise the secondary standards by adding secondary standards identical to the NO2 and SO2

18 primary 1-hour standards set in 2010. While this will not add secondary standards of the 

19 ecologically relevant form to address acid deposition, it will directionally provide some degree of 

20 additional protection. This is consistent with the view that the current secondary standards are 

21 neither sufficiently protective nor appropriate in form, but that it is not appropriate to propose to 

22 set a new, ecologically relevant multi-pollutant secondary standard at this time, for all of the 

23 reasons discussed above.
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1 While not a basis for this decision, the Administrator also recognizes that a new, 

2 innovative AAI-based standard would raise significant implementation issues that would need to 

3 be addressed consistent with the CAA requirements for implementation-related actions following 

4 the setting of a new NAAQS. It will take time to address these issues, during which the Agency

5 will be conducting a field pilot program to gather necessary data and the environment will 

6 benefit from reductions in oxides of nitrogen and sulfur resulting from the new NO2 and SO2

7 primary standards, as noted above, as well as reductions expected to be achieved from EPA’s 

8 proposed Transport Rule and Mercury and Air Toxics standards. These implementation-related 

9 issues are discussed in more detail below in section IV.A.5.

10 The Administrator solicits comment on all aspects of this proposed decision, including 

11 the framework and elements of a multi-pollutant standard for oxides of nitrogen and sulfur to 

12 address deposition-related effects on sensitive ecosystems, with a focus on aquatic acidification, 

13 and the uncertainties and complexities associated with the development of such a standard at this 

14 time. The Administrator also solicits comment on the field pilot program and related monitoring 

15 methods as discussed below in section IV.

16 IV. Field Pilot Program and Ambient Monitoring 

17 This section describes EPA’s plans for a field pilot program and the evaluation of 

18 monitoring methods for ambient air indicators of oxides of nitrogen (NOy) and oxides of sulfur 

19 (SOx) to implement the Administrator’s decision to undertake such a field monitoring program 

20 in conjunction with her decision to propose not to set a new multi-pollutant secondary standard 

21 in this review, as discussed above in section III.H. As noted above and discussed below in 

22 section IV.A, the field pilot program is intended to collect and analyze data so as to enhance our 

23 understanding of the degree of protectiveness that would likely be afforded by a standard based on 

24 the AAI as developed in the Policy Assessment. Data generated by this field program would also 
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1 support development of an appropriate monitoring network that would work in concert with such a 

2 standard to result in the intended degree of protection. As discussed below in section IV.B, the 

3 evaluation of monitoring methods focuses on the development of Federal Reference 

4 Methods/Federal Equivalent Methods (FRM/FEM) for NOy and SOx. 

5 A. Field Pilot Program

6 This section presents the objectives of a field pilot program (section IV.A.1) that would 

7 gather relevant field data over a five-year period in a sample of 3 to 5 sensitive ecoregions across 

8 the country. An overview of the scope and structure of the field program, with a focus on 

9 measurements of ambient air indicators of oxides of nitrogen and sulfur, is presented in section 

10 IV.A.2. Section IV.A.3 explains the role of additional complementary measurements beyond the 

11 ambient air indicators that would be included in the program, and section IV.A.4 discusses a 

12 parallel longer-term research agenda, both of which are guided by the uncertainties discussed 

13 above in section III. Section IV.A.5 identifies implementation challenges presented by an AAI14 based standard that could be addressed in parallel with a field pilot program. Section IV.A.6 

15 discusses engagement with stakeholder groups as part of the planned pilot program.

16 1. Objectives

17 Consideration of a new multi-pollutant standard to address deposition-related effects on 

18 sensitive aquatic ecoregions raises unique challenges relative to those typically raised in reviews 

19 of existing NAAQS for which an established network of FRM/FEM monitors, designed to 

20 measure the indicator pollutant, is generally available. The primary goal of this field pilot 

21 program, and the related monitoring program discussed in section IV.B, is to enhance our 

22 understanding of the degree of protectiveness that would likely be afforded by a standard based 

23 on the AAI, as described above in section III, and to support the identification of a specific form 
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1 and level for such a standard that would be requisite to protect public welfare consistent with the 

2 requirements of the CAA, through the following objectives:

3 (1) evaluate measurement methods for the ambient air indicators of NOy and SOx and 

4 consider designation of such methods as FRMs;

5 (2) examine the variability and improve characterization of concentration and 

6 deposition patterns of oxides of nitrogen and sulfur, as well as reduced forms of 

7 nitrogen, within and across a number of sensitive ecoregions across the country;

8 (3) develop updated ecoregion-specific factors (i.e., F1 through F4) for the AAI 

9 equation based in part on new observed air quality data within the sample ecoregions 

10 as well as on updated nationwide air quality model results and expanded critical load 

11 data bases, and explore alternative approaches for developing such representative 

12 factors;

13 (4) calculate ecoregion-specific AAI values using observed NOy and SOx data and 

14 updated ecoregion-specific factors to examine the extent to which the sample 

15 ecoregions would meet a set of alternative AAI-based standards;

16 (5) develop air monitoring network design criteria for an AAI-based standard;

17 (6) assess the use of total nitrate measurements as a potential alternative indicator for 

18 oxides of nitrogen; and

19 (7) support related longer-term research efforts, including enhancements to and 

20 evaluation of modeled dry deposition algorithms.

21 2. Overview of Field Pilot Program

22 The CASTNET program (Figure IV-1) affords an available infrastructure relevant to an 

23 AAI-based standard, given the location of sites in some acid-sensitive ecoregions and various 
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1 measurements of sulfur and nitrogen species. The EPA plans to use CASTNET sites in selected 

2 acid-sensitive ecoregions to serve as the platform for this pilot program, potentially starting in 

3 late 2012 and extending through 2018. CASTNET sites in three to five ecoregions in acid4 sensitive areas would collect NOy and SOx (i.e., SO2 and p-SO4) measurements over a five-year 

5 period. The initial step in developing a data base of observed ambient air indicators for oxides of 

6 nitrogen and sulfur requires the addition of NOy samplers at the pilot study sites so that a full 

7 complement of indicator measurements are available to calculate AAI values. These CASTNET 

8 sites would also be used to make supplemental observations useful for evaluation of CMAQ’s 

9 characterization of factors F2 –F4 in the AAI equation.

10 The selected ecoregions would account for geographic variability by including regions 

11 from across the U.S., including the east, upper midwest and west. Each selected region would 

12 have at least two existing CASTNET sites. Each of the pilot CASTNET sites would be used to 

13 verify and enhance the performance of the established methods, data retrieval and reporting 

14 procedures. This would in turn inform the future selection of specific parameters for the AAI 

15 equation.

16
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1

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

3 

4 Over the course of this five-year pilot program, the most current national air quality

5 modeling, based on the most current national emissions inventory, would be used to develop an 

6 updated set of F2 – F4 factors. A parallel multi-agency national critical load data base

7 development effort would be used as the basis for calculating updated F1 factors. As discussed 

8 above in section III.B, these factors would be based on average parameter values across an 

9 ecoregion. Using this new set of F factors, observations of NOy and SOx derived from the pilot 

10 program, averaged across each ecoregion, would be used to calculate AAI values in the sample 

11 ecoregions. The data from the pilot program would also be used to examine alternative 
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1 approaches to generating representative air quality values, such as examining the appropriateness 

2 of spatial averaging in areas of high spatial variability.

3 3 . Complementary Measurements

4 Complementary measurements may be performed at some sites in the pilot network to 

5 reduce uncertainties in the recommended methods and better characterize model performance 

6 and application to the AAI. The CASAC Air Monitoring and Methods Subcommittee (AMMS) 

7 advised EPA that such supplemental measurements were of critical importance in a field 

8 measurement program related to an AAI-based standard (Russell and Samet, 2011b).

9 Candidate complementary measurements to address sulfur, in addition to those provided 

10 by the CASTNET filter pack (CFP), include trace gas continuous SO2 and speciated PM2.5 

11 measurements. The co-located deployment of a continuous SO2 analyzer with the CFP for SO2

12 will provide test data for determining suitability of continuous SO2 measurements as an FEM, as 

13 well as producing valuable time series data for model evaluation purposes. The weekly 

14 averaging time provided by the CFP adequately addresses the annual-average basis of an AAI15 based secondary standard, but would not be applicable to short-term (i.e., 1-hour) averages

16 associated with the primary SO2 standard. Conversely, because of the low concentrations 

17 associated with many acid-sensitive ecoregions, existing SO2 FRMs designated for use in 

18 determining compliance with the primary standard would not necessarily be appropriate for use 

19 in conjunction with an AAI-based secondary standard. 

20 Co-locating the PM2.5 sampler used in the EPA Chemical Speciation Network and the 

21 IMPROVE network at pilot network sites would allow for characterizing the relationship 

22 between the CFP-derived p-SO4 and the speciation samplers used throughout the state and local 

23 air quality networks. Note that CASTNET already has several co-located IMPROVE chemical 
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1 speciation samplers. Because the AAI equation is based on concentration of p-SO4, the original 

2 motivation for capturing all particle size fractions is not as important relative to simply capturing 

3 the concentration of total p-SO4. 

4 Candidate measurements to complement oxidized nitrogen measurements, in addition to 

5 the CFP, include a mix of continuous and periodic sampling for the dominant NOy species, 

6 namely NO, true NO2, PAN, HNO3, and p-NO3. While there are several approaches to acquiring 

7 these measurements, perhaps the most efficient strategy would take advantage of the available 

8 CFP for total nitrate, and add a three-channel chemiluminescence instrument that will cycle 

9 between NOy, true NO2 and NO by adding photolytic detection for true NO2. Other options for 

10 measuring true NO2 would include adding either a stand-alone photolytic or cavity ring-down 

11 spectroscopy instrument. Measurements of PAN may be acquired either on a periodic basis 

12 through canister sampling and subsequent laboratory analysis or through emerging in-situ 

13 sampling and analysis methods. Although the CFP yields a reliable measurement of total nitrate, 

14 the t-NO3 (i.e., the sum of HNO3 and p-NH4) value, strong consideration may be given to direct 

15 measurement of HNO3, which has the highest deposition velocity of all the dominant NOy

16 species. Similar to the use of continuous SO2 data, these speciated NOy data serve two purposes:

17 evaluating total NOy instrument behavior, and air quality model evaluation. The measurement 

18 of individual NOy species can be used to generate site-specific NOy values for comparison to 

19 modeled NOy, and will likely provide insight into and improvement of modeled dry deposition.

20 The CASAC AMMS (Russell and Samet, 2011b) recommended that EPA consider the 

21 use of t-NO3 obtained from CASTNET sampling as an indicator for NOy, reasoning that t-NO3 is 

22 typically a significant fraction of deposited oxidized nitrogen in rural environments and 

23 CASTNET measurements are widely available. Collection of this data would support further
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1 consideration of using the CFP for t-NO3 as the indicator of oxides of nitrogen for use in an AAI2 based secondary standard. 

3 The CASAC AMMS also recommended that total NHx (NH3 and p-NH4) be considered 

4 as a proxy for reduced nitrogen species, reasoning that the subsequent partitioning to NH3 and p5 NH4 may be estimated using equilibrium chemistry calculations. Reduced nitrogen 

6 measurements are used to evaluate air quality modeling which is used in generating factor F2. 

7 Additional studies are needed to determine the applicability of NHx measurements and 

8 calculated values of NH3 and NH4 to the AAI. 

9 The additional supplemental measurements of speciated NOy, continuous SO2 and NHx 

10 will be used in future air quality modeling evaluation efforts. Because there often is significant 

11 lag in the availability of contemporary emissions data to drive air quality modeling, the complete 

12 use of these data sets will extend beyond the five-year collection period of the pilot program. 

13 Consequently, the immediate application of those data will address instrument performance 

14 comparisons that explore the feasibility of using continuous SO2 instruments in rural 

15 environments, and using the speciated NOy data to assess NOy instrument performance. 

16 Although contemporary air quality modeling will lag behind measurement data availability, the 

17 observations can be used in deposition models to compare observed transference ratios with the 

18 previously calculated transference ratios to test temporal stability of the ratios. 

19 An extended water quality sampling effort should parallel the air quality measurement 

20 program to address some of the uncertainties related to factor F1 and the representativeness of 

the nth 21 percentile critical load as discussed in section III.B.5.b.i. The objective of the water 

22 quality sampling would be to develop a larger data base of critical loads in each of the pilot 

ecoregions such that the nth 23 percentile can adequately be characterized in terms of representing 
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1 all water bodies. Opportunities to leverage and perhaps enhance existing ecosystem modeling 

2 efforts enabling more advanced critical load modeling and improved methods to estimate base 

3 cation production would be pursued. For example, areas with ongoing research studies 

4 producing data for dynamic critical load modeling would be considered when selecting the pilot 

5 ecoregions.

6 4. Complementary areas of research 

7 The EPA recognizes that a source of uncertainty in an AAI-based secondary standard that 

8 would not be directly addressed in the pilot program stems from the uncertainty in the model 

9 used to link atmospheric concentrations to dry deposition fluxes. Currently, there are no ongoing

10 direct dry deposition measurement studies at CASTNET sites that can be used to evaluate 

11 modeled results. It was strongly recommended by CASAC AMMS that a comprehensive 

12 sampling-intensive study be conducted in at least one, preferably two sites in different 

13 ecoregions to assess characterization of dry deposition of sulfur and nitrogen. These sites would 

14 be the same as those for the complementary measurements described above, but they would

15 afford an opportunity to also complement dry deposition process research that benefits from the 

16 ambient air measurements collected in the pilot program. The concerns regarding uncertainties 

17 underlying an AAI-based secondary standard suggest that research that includes dry deposition 

18 measurements and evaluation of dry deposition models should be a high priority.

19 Similar leveraging should be pursued with respect to ecosystem research activities. For 

20 example, studies that capture a suite of soil, vegetation, hydrological, and water quality 

21 properties that can help evaluate more advanced critical load models would complement the 

22 atmospheric-based pilot program. In concept, such studies could provide the infrastructure for 

23 true multi-pollutant, multi-media “super” sites assuming the planning, coordination, and resource 
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1 facets can be aligned. While this discussion emphasizes the opportunity of leveraging ongoing 

2 research efforts, consideration could be given to explicitly including related research components 

3 directly in the pilot program.

4 5. Implementation challenges

5 The CAA requires that once a NAAQS is established, designation and implementation 

6 must move forward. With a standard as innovative as the AAI-based standard considered in this 

7 review, the Administrator believes that its success will be greatly improved if, while additional 

8 data are being collected to reduce the uncertainties discussed above, the implementing agencies 

9 and other stakeholders have an opportunity to discuss and thoroughly understand how such a 

10 standard would work. And since, as noted above, emissions reductions that are directionally

11 correct to reduce aquatic acidification will be occurring as a result of other CAA programs, the 

12 Administrator believes that this period of further discussion will not delay progress but will 

13 ensure that once implementation is triggered, agencies will be prepared to implement it 

14 successfully.

15 Consideration of an AAI-based secondary standard for oxides of nitrogen and sulfur 

16 would present significant implementation challenges because it involves multiple, regionally17 dispersed pollutants and relatively complex compliance determinations. The anticipated 

18 implementation challenges fall into three main categories: monitoring and compliance 

19 determinations for area designations, pre-construction permit application analyses of individual 

20 source impacts, and State Implementation Plan (SIP) development. Several overarching 

21 implementation questions that we anticipate will be addressed in parallel with the field pilot 

22 program’s five-year data collection period include: 

23 (1) What are the appropriate monitoring network density and siting requirements to 
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1 support a compliance system based on ecoregions? 

2 (2) Given the unique spatial nature of the secondary standard (e.g., ecoregions), what are 

3 the appropriate parameters for establishing nonattainment areas?

4 (3) How can new or modified major sources of oxides of nitrogen and oxides of sulfur 

5 emissions assess their ambient impacts on the standard and demonstrate that they are not 

6 causing or contributing to a violation of the NAAQS for preconstruction permitting?

7 (4) What additional tools, information, and planning structures are needed to assist states 

8 with SIP development, including the assessment of interstate pollutant transport and 

9 deposition?

10 (5) Would transportation conformity apply in nonattainment and maintenance areas for 

11 this secondary standard, and, if it does, would satisfying requirements that apply for 

12 related primary standards (e.g., ozone, PM2.5, and NO2) be demonstrated to satisfy

13 requirements for this secondary standard?

14 6. Final monitoring plan development and stakeholder participation

15 The existing CASTNET sampling site infrastructure provides an effective means of 

16 quickly and efficiently deploying a monitoring program to support potential implementation of 

17 an AAI-based secondary standard, and also provides an additional opportunity for federally 

18 managed networks to collaborate and support the states, local agencies and Tribes (SLT) in 

19 determining compliance with a secondary standard. A collaborative effort would help to 

20 optimize limited federal and SLT monitoring funds and would be beneficial to all involved. 

21 CASTNET is already a stakeholder-based program with over 20 participants and contributors, 

22 including federal, state and tribal partners. CASTNET partners are willing to adapt the network 
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1 to meet the monitoring requirements of an AAI-based secondary standard, provided the current 

2 requirements for program assessment continue to be met. 

3 The CASAC AMMS generally endorsed the technical approaches used in CASTNET, but 

4 concerns were raised by individual representatives of State agencies concerning the perception of 

5 EPA-controlled management aspects of CASTNET and data ownership. Potential approaches to 

6 resolve these issues will be developed and evaluated in existing NACAA/EPA ambient air 

7 monitoring workgroups. The EPA Office of Air and Radiation (which includes the Office of Air 

8 Quality Planning Standards, OAQPS; and the Office of Atmospheric Program’s Clean Air 

9 Markets Division, OAP-CAMD), and their partners on the NACAA monitor steering committee 

10 will develop a prioritized specific plan that identifies the 3-5 ecoregions and the instrumentation 

11 to be deployed. EPA anticipates that a cost estimate of the plan with priorities and options will 

12 be developed by January, 2012. Although this pilot program is focused on data collection, the 

13 plan will include details of the data analysis approaches as well as a vehicle that incorporates 

14 engagement from those within EPA and SLTs to foster progress on the implementation questions 

15 noted above in section IV.A.5.

16 If an AAI-based secondary standard were to be set in the future, deployment of a full 

17 national network would follow the pilot monitoring program. The number of sites deployed in 

18 the network will lead to increased confidence in capturing spatial patterns of air quality. 

19 Recognizing that this section presents the general elements of the field pilot programs, EPA 

20 intends to develop a more detailed field pilot program plan through a process that will engage the 

21 air quality management and research (atmospheric and ecosystem) communities, as well as other 

22 federal agencies, state and local agencies, and non-government based centers of expertise. The 

23 EPA is seeking comment and input on all aspects of this field pilot program.
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1 .B. Evaluation of Monitoring Methods

2 The EPA generally relies on monitoring methods that have been designated as FRMs or 

3 FEMs for the purpose of determining the attainment status of areas with regard to existing 

4 NAAQS. Such FRMs or FEMs are generally required to measure the air quality indicators that 

5 are compared to the level of a standard to assess compliance with a NAAQS. Prior to their 

6 designation by EPA as FRM/FEMs through a rulemaking process, these methods must be 

7 determined to be applicable for routine field use and need to have been experimentally validated 

8 by meeting or exceeding specific accuracy, reproducibility, and reliability criteria established by 

9 EPA for this purpose. As discussed above in section III.A, the ambient air indicators being 

10 considered for use in an AAI-based standard include SO2, particulate sulfate (p-SO4), and total 

11 reactive oxides of nitrogen (NOy).

12 The CASTNET provides a well established infrastructure that would meet the basic 

13 location and measurement requirements of an AAI-based secondary standard given the rural 

14 placement of sites in acid sensitive areas. In addition, CFPs currently provide very economical 

15 weekly, integrated average concentration measurements of SO2, p-SO4, ammonium ion (NH4) 

16 and t-NO3, the sum of HNO3 and p-NO3. 

17 While routinely operated instruments that measure SO2, p-SO4, NOy and/or t-NO3 exist, 

18 instruments that measure p-SO4, NOy, t-NO3, or the CFP for SO2 have not been designated by 

19 EPA as FRMs or FEMs The EPA’s ORD has initiated work that will support future FRM 

20 designations by EPA for SO2 and p-SO4 measurements based on the CFP. Such a designation by

21 EPA could be done for the purpose of facilitating consistent research related to an AAI-based 

22 standard and/or in conjunction with setting and supporting an AAI-based secondary standard. 
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1 Based on extensive review of literature and available data, the EPA has identified 

2 potential methods that appear suitable for measuring each of the three components of the 

3 indicators. These three methods are being considered as new FRMs to be used for measuring the 

4 ambient concentrations of the three components that would be needed to determine compliance 

5 with an AAI-based secondary standard.

6 For the SO2 and p-SO4 measurements, EPA is considering the CFP method, which 

7 provides weekly average concentration measurements for SO2 and p-SO4. This method has been 

8 used in the EPA’s CASTNET monitoring network for 15 years, and strongly indicates that it will 

9 meet the requirements for use as an FRM for the SO2 and p-SO4 concentrations for an AAI-based 

10 secondary standard. 

11 Although the CFP method would provide measurements of both the SO2 and p-SO4 

12 components in a unified sampling and analysis procedure, individual FRMs will be considered 

13 for each. The EPA recognizes that an existing FRM to measure SO2 concentrations using ultra14 violet fluorescence (UVF) exists (40 CFR Part 50, Appendix A-1) for the purpose of monitoring 

15 compliance for the primary SO2 NAAQS. However, several factors suggest that the CFP method 

16 would be superior to that UVF FRM for monitoring compliance with an AAI-based secondary

17 standard and will be discussed in more detail below. 

18 For monitoring the NOy component, a continuous analyzer for measuring NOy is 

19 commercially available and is considered to be suitable for use as an FRM. This method is 

20 similar in design to the existing NO2 FRM (described in 40 CFR Part 50, Appendix F), which is 

21 based on the ozone chemiluminescence measurement technique. The method is adapted to and 

22 further optimized to measure all oxides of nitrogen (NOy). However, this NOy method requires 

23 further evaluation before it can be fully confirmed as a suitable FRM. The EPA is currently 
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1 completing a full scientific assessment of the NOy method to determine whether it would be 

2 appropriate to consider for designation by EPA as an FRM. Specific details on these three 

3 methods are given below. 

4 On February 16, 2011, EPA presented this set of potential FRMs to the CASAC AMMS 

5 for their consideration and comment. In response, the CASAC AMMS stated that, overall, it 

6 believes that EPA’s planned evaluation of methods for measuring NOy, SO2 and p-SO4 as 

7 ambient air indicators are suitable approaches in concept. On supporting the CFP method as a 

8 potential FRM for SO2, CASAC stated that they felt that the CFP is adequate for measuring 

9 long-term average SO2 gas concentrations in rural areas with low levels (less than 5 ppbv) and is 

10 therefore suitable for consideration as an FRM. For p-SO4, CASAC generally supports the use 

11 of the CFP as a potential FRM for measuring p-SO4 for an AAI-based secondary standard. The 

12 method has been relatively well-characterized and evaluated, and it has a documented, long-term 

13 track record of successful use in a field network designed to assess spatial patterns and long-term 

14 trends.

15 On supporting the photometric NOy method as a potential FRM, CASAC concluded that 

16 the existing NOy method is generally an appropriate approach for the indicator. However, 

17 CASAC agrees that additional characterization and research is needed to fully understand the 

18 method in order to designate it as a FRM. The EPA is now soliciting public comment on these 

19 methods as to their adequacy, suitability, and relative merits as FRMs for purposes of monitoring 

20 to determine compliance with an AAI-based secondary standard.

21 1. Potential FRMs for SO2 and p-SO4

22 The CFP is a combined, integrated sampling and analysis method based on the well23 established measurement technology that has been used extensively in EPA’s CASTNET 
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1 monitoring network (see http://www.epa.gov/castnet). This method is in current use at over 80 

2 monitoring sites and has been in use at not less than 40 sites for over 15 years. This method 

3 employs a relatively simple and inexpensive sampler and uses four 47-mm filters placed in an 

4 open-faced filter pack to simultaneously collect integrated filter samples for the SO2 and p-SO4

5 components. In addition, the CFP is also capable of the collection of t-NO3, the sum of HNO3

6 and p-NO3.

The first stage of the filter pack assembly contains a Teflon® filter that collects p-SO4

2-

7

and p-NO3, the second stage contains a nylon filter that collects SO2 (as SO4

2-

8 ) and HNO3, and 

9 the third stage contains two cellulose fiber filters impregnated with potassium carbonate (K2CO3) 

that collect any remaining SO2 (as SO4

2-

10 ). The sampler collects one-week integrated samples at a

11 very low, controlled flow rate (1.5 or 3 L/min) in an attempt simulate actual deposition. Weekly

12 averaged SO2 and p-SO4 concentrations could then be averaged over a 1-year period to calculate 

13 annual average values.

14 Upon sample completion, the species-specific filters are extracted, with subsequent 

15 analysis by the well-established and documented ion chromatographic (IC) analytical technique. 

16 During the IC analysis, an aliquot of a filter extract is injected into a stream of eluent (ion 

17 chromatography mobile phase, generally a millimolar-strength solution of carbonate18 bicarbonate) and passed through a series of ion exchangers. The anions of interest are separated 

19 on the basis of their relative affinities for a low capacity and the strongly basic anion exchanger 

20 (guard and separator column). The separated anions are directed onto a cation exchanger 

21 (suppressor column) where they are converted to their highly conductive acid form, and the 

22 eluent is converted to a weakly conductive form. The now-separated anions, each in their acid 

23 form, are measured by conductivity. They are identified on the basis of retention time compared 
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1 to that of standards and quantified by measurement of peak area compared to the peak areas of 

2 calibration standards. 

3 Calibration and quality assurance for the method are applied to the sample filters, the 

4 analytical processes, and the flow rate measurement and control aspects of the sampler. Overall 

5 method performance is typically assessed with collocated samplers. These quality assurance 

6 techniques are routinely used and have proved adequate for other types of FRMs and equivalent 

7 methods in air monitoring network service.

8 The measurement and analytical procedures and past performance data associated with 

9 the CFP method are well documented and available through Quality Assurance Performance 

10 Plans (QAPPs), Standard Operating Procedures (SOPs) and annual reports (US EPA, 2010a and 

11 2010b). The accumulated database on the CFP method is substantial and indicates that the 

12 method is sound, stable and has good reliability in routine, field operation. Data quality 

13 assessment results show the method to have good reproducibility, with collocated and analytical 

14 precision values in the range of 2% – 10% (excluding very low concentration measurements near 

15 the method detection limits; US EPA 2010b).

16 Data quality objectives (DQOs) for a new FRM would be based upon current DQOs 

17 being used for this method by EPA’s OAP/CAMD and the National Park Service (NPS), the 

18 federal managers of CASTNET (US EPA, 2010a). In its current state, the CFP method is 

19 expected to meet or exceed (as past CASTNET data have indicated; US EPA, 2010b) the 

20 expected FRM DQOs, even when deployed in new monitoring networks outside of CASTNET. 

21 In addition, CASTNET samples have agreed favorably with other measures of SO2 and p-SO4 in 

22 comparison studies. For example, in direct comparison with an annular denuder sampler (ADS) 

23 method, CASTNET/ADS ratios for SO2 and p-SO4 were generally on the order of 0.9-1.1 
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1 (Lavery et al, 2009; Sickles et al, 1999; Sickles et al, 2008), thus illustrating the accuracy of the 

2 CFP method in the determination of long-term average SO2 and p-SO4 concentrations. The EPA 

3 believes that the CFP method would be fully adequate as an FRM in determining yearly average 

4 SO2 and p-SO4 concentrations for compliance determination purposes.

5 The EPA recognizes that an existing FRM for SO2 has proven adequate for the purposes 

6 of monitoring compliance for the primary SO2 NAAQS, specifically the newly-promulgated 1-

7 hour standard. However, this FRM is better suited to the shorter-term, higher concentration 

8 primary and secondary SO2 NAAQS, and there is substantial uncertainty as to the adequacy of 

9 this SO2 FRM for monitoring the lower concentrations relevant to determining compliance with 

10 an AAI-based secondary standard. The performance specifications for SO2 FRM analyzers (40 

11 CFR Part 53, Table B-1) require a lower detectable limit (LDL) of 0.002 ppm for the standard 

12 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/m313 , respectively. 

14 Analysis of 2009 CASTNET data shows that of the 84 CASTNET sampling sites, 63 measured 

15 annual average SO2 concentrations below even the lower of these LDL requirements of 2.62 

μg/m316 for the lower range SO2 FRM (US EPA, 2010a). In addition, 11 of the 84 sites measured 

17 annual (2009) average SO2 concentrations very near or below the manufacturers’ reported 

18 detection limits for trace level UVF SO2 monitors. Further, it is likely that the number of sites 

19 with annual average SO2 concentration below both the SO2 FRM LDL and the manufacturers 

20 reported detection limits will increase due to expected declines in mean SO2 concentrations (US

21 EPA, 2010b). For these reasons, EPA is considering the CFP method for use as the FRM for 

22 monitoring the SO2 component of an ambient air indicator for oxides of sulfur, with a 
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1 recommendation for additional study and data collection to further evaluate the possible 

2 applicability of the continuous UVF SO2 FRM for this purpose.

3 2. Potential FRM for NOy

4 Atmospheric concentrations of NOy are measured continuously by an analyzer that 

5 photometrically measures the light intensity, at wavelengths greater than 600 nanometers (nm), 

6 resulting from the chemiluminescent reaction of ozone (O3) with NO in sampled air. This 

7 method is very similar to the chemiluminescence NO/NO2 analyzers widely used to collect NO2

8 monitoring data for determining compliance with the NO2 NAAQS. The various oxides of 

9 nitrogen species, excluding NO, are first quantitatively reduced to NO by means of a catalytic 

10 converter. These species include NO2, HNO2, PANs, HNO3 and p-NO3. The NO, which 

11 commonly exists in ambient air, passes through the converter unchanged, and, when combined 

12 with the NO resulting from the catalytic conversion of the other oxides of nitrogen, a 

13 measurement of the total NOy concentration results. To maximize the conversion of the more 

14 chemically active oxides of nitrogen species, the converter is located externally, at or near the air 

15 sample inlet probe. This location minimizes losses of these active species that could otherwise 

16 occur from chemical reactions and wall losses in the sample inlet line.

17 The NOy analyzer is a suitable, commercially produced continuous chemiluminescence 

18 analyzer that includes an ozone generator, a reaction cell, a photometric detector, wavelength 

19 filters as necessary to reduce sensitivity to wavelengths below 600 nm, a pump and flow control 

20 system to draw atmospheric air through the converter and into the reaction cell, a suitable 

21 converter, a system to control the operation of the analyzer, and appropriate electronics to 

22 process and quantitatively scale the photometric signals. The converter contains a catalyst such 

23 as molybdenum and is heated to an optimum temperature designed to optimize the conversion of 
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1 the various oxides of nitrogen to NO. It is connected to the analyzer via suitable lengths of 

2 Teflon® tubing. Hourly NOy measurements obtained by the analyzer would be averaged over 

3 the same 7-day period used by the CFP method to measure the SO2 and p-SO4 components, with 

4 further averaging over a 1-year period.

5 Commercial NOy analyzers are currently available, and the analyzers have been used for 

6 a variety of monitoring applications. During the 2006 TexAQS Radical and Aerosol 

7 Measurement Project (TRAMP), Luke et. al., 2010, compared measured NOy concentrations 

8 obtained with an NOy instrument based upon the above mentioned methodology with the sum of 

9 measured individual NOy species (i.e., NOyi = NO+NO2+HNO3+PANs+HNO2+p-NO3). This 

10 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; r211 = 0.9975) 

12 periods (Luke et al, 2010). The results of this study show that this NOy method is capable of the 

13 accurate determination of all the atmospherically relevant NOy components, resulting in an 

14 accurate determination of total NOy concentrations. NOy instruments have been routinely 

15 operated in networks such as SEARCH, dating back several years. In addition, state monitoring 

16 agencies across the U.S. have begun, starting in 2009, the routine operation of commercially

17 available NOy instrumentation in anticipation of EPA’s NCore network transitioning to full 

18 operation in 2011. 

19 These initial assessments described above are promising and indicate that the photometric 

20 NOy method appears to be accurate, reliable, and capable of routine network operation. As a

21 result, the method is most likely capable for use as an FRM for determining atmospheric NOy

22 concentrations as a component in determining compliance with an AAI-based secondary

23 standard. However, as described below, this continuous method for NOy requires additional 
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1 time for further evaluation before it can be fully confirmed for adoption as a FRM. The EPA has 

2 identified measurement uncertainties and some remaining science questions associated with this 

3 method. Among these are (a) the ability of the method to capture all components of NOy

4 relevant to nitrogen deposition, (b) the efficiency of the molybdenum converter in converting all 

5 oxides of nitrogen to NO for detection (excluding NO2, as this conversion is already well 

6 documented), (c) appropriate inlet height specifications to minimize any bias associated with 

7 vertical concentration gradients of key NOy components, (d) identification and quantification of 

8 potential measurement interferences in the NOy determination, and (e) development and 

9 demonstration of effective calibration/challenge procedures to best represent the various 

10 mixtures of NOy components that are expected to be present in the different air sheds across the 

11 U.S.

12 To address these NOy method uncertainties and to fully assess this method for use as the 

13 NOy FRM, EPA has developed a detailed research plan (Russell and Samet, 2011b) which was 

14 presented to the CASAC AMMS on February 16, 2011. In response, CASAC recognized the 

15 need for, and supported the general outline of EPA’s research plan to evaluate the NOy method 

16 for potential designation as an FRM (US EPA, 2011). In addition, the CASAC AMMS 

17 suggested additional areas of research associated with the photometric NOy method that warrant 

18 further assessment prior to final designation of the method as the NOy FRM. These include 

19 operation of the method during extremely low temperature conditions to investigate possible 

20 condensation in sample lines, method detection limits relative to low levels expected in remote 

21 areas, and ambient-based method evaluations in various air sheds across the U.S. In response to 

22 these CASAC AMMS suggestions, EPA is carrying out studies, in addition to the tasks outlined 

23 in the research plan, for the NOy method. The results of these studies will likely take a year or 
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1 more to become available. As noted previously, EPA anticipates that these results will be 

2 favorable and will confirm the adequacy of the NOy method as a suitable FRM for determining 

3 compliance with an AAI-based secondary standard.

4 V. Statutory and Executive Order Reviews

5 A. Executive Order 12866: Regulatory Planning and Review and Review and Executive 

6 Order 13563: Improving Regulation and Regulatory Review

7 Under Executive Order 12866 (58 FR 51735, October 4, 1993), this action is a

8 “significant regulatory action.” Accordingly, EPA submitted this action to the Office of 

9 Management and Budget (OMB) for review under Executive Orders 12866 and 13563 (76 FR 

10 3821, January 21, 2011), and any changes made in response to OMB recommendations have 

11 been documented in the docket for this action.

12 B. Paperwork Reduction Act

13 This action does not impose an information collection burden under the provisions of the 

14 Paperwork Reduction Act, 44 U.S.C. 3501 et seq. Burden is defined at 5 CFR 1320.3(b). There 

15 are no information collection requirements directly associated with the establishment of a 

16 NAAQS under section 109 of the CAA.

17 C. Regulatory Flexibility Act

18 For purposes of assessing the impacts of today’s rule on small entities, small entity is 

19 defined as: (1) a small business that is a small industrial entity as defined by the Small Business 

20 Administration’s (SBA) regulations at 13 CFR 121.201; (2) a small governmental jurisdiction 

21 that is a government of a city, county, town, school district or special district with a population of 

22 less than 50,000; and (3) a small organization that is any not-for-profit enterprise which is 

23 independently owned and operated and is not dominant in its field.
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1 After considering the economic impacts of today's proposed rule on small entities, I

2 certify that this action will not have a significant economic impact on a substantial number of 

3 small entities. This proposed rule will not impose any requirements on small entities. Rather, 

4 this rule establishes national standards for allowable concentrations of oxides of nitrogen and 

5 sulfur in ambient air as required by section 109 of the CAA. See also American Trucking 

6 Associations v. EPA. 175 F. 3d at 1044-45 (NAAQS do not have significant impacts upon small 

7 entities because NAAQS themselves impose no regulations upon small entities). We continue to 

8 be interested in the potential impacts of the proposed rule on small entities and welcome 

9 comments on issues related to such impacts

10 D. Unfunded Mandates Reform Act

11 Title II of the Unfunded Mandates Reform Act of 1995 (UMRA), Public Law 104-4, 

12 establishes requirements for Federal agencies to assess the effects of their regulatory actions on 

13 State, local, and Tribal governments and the private sector. Under section 202 of the UMRA, 

14 EPA generally must prepare a written statement, including a cost-benefit analysis, for proposed 

15 and final rules with “Federal mandates” that may result in expenditures to State, local, and Tribal 

16 governments, in the aggregate, or to the private sector, of $100 million or more in any 1 year. 

17 Before promulgating an EPA rule for which a written statement is needed, section 205 of the 

18 UMRA generally requires EPA to identify and consider a reasonable number of regulatory

19 alternatives and to adopt the least costly, most cost-effective or least burdensome alternative that 

20 achieves the objectives of the rule. The provisions of section 205 do not apply when they are 

21 inconsistent with applicable law. Moreover, section 205 allows EPA to adopt an alternative 

22 other than the least costly, most cost-effective or least burdensome alternative if the 

23 Administrator publishes with the final rule an explanation why that alternative was not adopted. 
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1 Before EPA establishes any regulatory requirements that may significantly or uniquely affect 

2 small governments, including Tribal governments, it must have developed under section 203 of 

3 the UMRA a small government agency plan. The plan must provide for notifying potentially 

4 affected small governments, enabling officials of affected small governments to have meaningful 

5 and timely input in the development of EPA regulatory proposals with significant Federal 

6 intergovernmental mandates, and informing, educating, and advising small governments on 

7 compliance with the regulatory requirements. 

8 This action contains no Federal mandates under the provisions of Title II of the 

9 Unfundated Mandates Reform Act of 1995 (UMRA), 2 U.S.C. 1531-1538 for State, local, or 

10 tribal governments or the private sector. Therefore, this action is not subject to the requirements 

11 of sections 202 or 205. Furthermore, as indicated previously, in setting a NAAQS EPA cannot 

12 consider the economic or technological feasibility of attaining ambient air quality standards; 

13 although such factors may be considered to a degree in the development of State plans to 

14 implement the standards. See also American Trucking Associations v. EPA, 175 F. 3d at 1043 

15 (noting that because EPA is precluded from considering costs of implementation in establishing 

16 NAAQS, preparation of a Regulatory Impact Analysis pursuant to the Unfunded Mandates 

17 Reform Act would not furnish any information which the court could consider in reviewing the 

18 NAAQS). Accordingly, EPA has determined that the provisions of sections 202, 203, and 205 of 

19 the UMRA do not apply to this proposed decision. The EPA acknowledges, however, that any

20 corresponding revisions to associated SIP requirements and air quality surveillance requirements, 

21 40 CFR part 51 and 40 CFR part 58, respectively, might result in such effects. Accordingly, 

22 EPA will address, as appropriate, unfunded mandates if and when it proposes any revisions to 40 

23 CFR parts 51 or 58.
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1 E. Executive Order 13132: Federalism

2 This proposed rule does not have federalism implications. It will not have substantial 

3 direct effects on the States, on the relationship between the national government and the States, 

4 or on the distribution of power and responsibilities among the various levels of government, as 

5 specified in Executive Order 13132 because it does not contain legally binding requirements. 

6 Thus, the requirements of Executive Order 13132 do not apply to this rule.

7 EPA believes, however, that this proposed rule may be of significant interest to State 

8 governments. As also noted in section E (above) on UMRA, EPA recognizes that States will 

9 have a substantial interest in this rule and any corresponding revisions to associated SIP 

10 requirements and air quality surveillance requirements, 40 CFR part 51 and 40 CFR part 58, 

11 respectively. Therefore, in the spirit of Executive Order 13132 and consistent with EPA policy 

12 to promote communications between EPA and State and local governments, EPA specifically 

13 solicits comment on this proposed rule from State and local officials.

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

15 Executive Order 13175, entitled “Consultation and Coordination with Indian Tribal 

16 Governments” (65 FR 67249, November 9, 2000), requires EPA to develop an accountable 

17 process to ensure “meaningful and timely input by tribal officials in the development of 

18 regulatory policies that have tribal implications.” This rule concerns the establishment of 

19 national standards to address the public welfare effects of oxides of nitrogen and sulfur.

20 This action does not have Tribal implications, as specified in Executive Order 13175 (65 

21 FR 67249, November 9, 2000). It does not have a substantial direct effect on one or more Indian 

22 tribes, since tribes are not obligated to adopt or implement any NAAQS. Thus, Executive Order 

23 13175 does not apply to this rule.
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1 G. Executive Order 13045: Protection of Children from Environmental Health & Safety 

2 Risks

3 This action is not subject to EO 13045 because it is not an economically significant rule 

4 as defined in EO 12866. 

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

6 Use

7 This action is not a “significant energy action” as defined in Executive Order 13211 (66 

8 FR 28355, May 22, 2001), because it is not likely to have a significant adverse effect on the 

9 supply, distribution, or use of energy. This action concerns the establishment of national 

10 standards to address the public welfare effects of oxides of nitrogen and sulfur. This action does 

11 not prescribe specific pollution control strategies by which these ambient standards will be met. 

12 Such strategies will be developed by States on a case-by-case basis, and EPA cannot predict 

13 whether the control options selected by States will include regulations on energy suppliers, 

14 distributors, or users.

15 I. National Technology Transfer and Advancement Act

16 Section 12(d) of the National Technology Transfer and Advancement Act of 1995 

17 (NTTAA), Public Law No. 104-113, 12(d) (15 U.S.C. 272 note) directs EPA to use voluntary

18 consensus standards in its regulatory activities unless to do so would be inconsistent with 

19 applicable law or otherwise impractical. Voluntary consensus standards are technical standards 

20 (e.g., materials specifications, test methods, sampling procedures, and business practices) that are 

21 developed or adopted by voluntary consensus standards bodies. The NTTAA directs EPA to 

22 provide Congress, through OMB, explanations when the Agency decides not to use available and 

23 applicable voluntary consensus standards.
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1 EPA is not aware of any voluntary consensus standards that are relevant to the 

2 provisions of this proposed rule. EPA welcomes any feedback on such standards that may be 

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

4 Minority Populations and Low-Income Populations

5 Executive Order 12898 (59 FR 7629 (Feb. 16, 1994)) establishes federal executive policy

6 on environmental justice. Its main provision directs federal agencies, to the greatest extent 

7 practicable and permitted by law, to make environmental justice part of their mission by

8 identifying and addressing, as appropriate, disproportionately high and adverse human health or 

9 environmental effects of their programs, policies, and activities on minority populations and low10 income populations in the United States. 

11 EPA has determined that this proposed rule will not have disproportionately high and 

12 adverse human health or environmental effects on minority or low-income populations because it 

13 retains the level of environmental protection for all affected populations without having any 

14 disproportionately high and adverse human health or environmental effects on any population, 

15 including any minority or low-income population.

16

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7 and Technical Information. OAQPS Staff Paper. EPA-450/5-82-007. U.S. Environmental Protection 

8 Agency, Office of Air Quality Planning and Standards, Research Triangle Park, NC.

9 US EPA, 1984a. The Acidic Deposition Phenomenon and Its Effects: Critical Assessment Review Papers. Volume I 

10 Atmospheric Sciences. EPA-600/8-83-016AF. Office of Research and Development, Washington, DC

11 US EPA, 1984b. The Acidic Deposition Phenomenon and Its Effects: Critical Assessment Review Papers. Volume 

12 II Effects Sciences. EPA-600/8-83-016BF. Office of Research and Development, Washington, DC

13 US EPA, 1985. The Acidic Deposition Phenomenon and Its Effects: Critical Assessment Document. EPA-600/8-

14 85/001. Office of Research and Development, Washington, DC

15 US EPA. 1995a. Review of the National Ambient Air Quality Standards for Nitrogen Dioxide: Assessment of 

16 Scientific and Technical Information. OAQPS Staff Paper. EPA-452/R-95-005. U.S. Environmental 

17 Protection Agency, Office of Air Quality Planning and Standards, Research Triangle Park, NC. September.

18 US EPA. 1995b. Acid Deposition Standard Feasibility Study Report to Congress. U.S. Environmental Protection 

19 Agency, Washington, DC. EPA-430/R-95-001a.

20 US EPA 2007. Integrated Review Plan for the Secondary National Ambient Air Quality Standards for Nitrogen 

21 Dioxide and Sulfur Dioxide. U.S. Environmental Protection Agency, Research Triangle Park, NC, EPA22 452/R-08-006.

23 US EPA 2008. Integrated Science Assessment (ISA) for Oxides of Nitrogen and Sulfur Ecological Criteria (Final 

24 Report). U.S. Environmental Protection Agency, Washington, D.C., EPA/600/R-08/082F, 2008.

25 US EPA 2009. Risk and Exposure Assessment for Review of the Secondary National Ambient Air Quality 

26 Standards for Oxides of Nitrogen and Oxides of Sulfur-Main Content - Final Report. U.S. Environmental 

27 Protection Agency, Washington, D.C., EPA-452/R-09-008a

28 US EPA, 2010a. CASTNET Quality Assurance Project Plan, Revision 7.0, October 2010, 

29 http://java.epa.gov/castnet/

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

31 US EPA 2011. Policy Assessment for the Review of the Secondary National Ambient Air Quality Standards for 

32 Oxides of Nitrogen and Oxides of Sulfur. U.S. Environmental Protection Agency, Washington, D.C., EPA33 452/R-11-005a.

34 US EPA, 2011b. Federal Reference Methods for NOy and p-SO4 for the New Combined NOx and SOx Secondary 

35 NAAQS Research Plan, EPA/600/1-11/002 January 20, 2011.

36 Wolff, G. T. 1993. CASAC closure letter for the 1993 Criteria Document for Oxides of Nitrogen addressed to U.S. 

37 EPA Administrator Carol M. Browner dated September 30, 1993.

38 Wolff, G. T. 1995. CASAC closure letter for the 1995 OAQPS Staff Paper addressed to U.S. EPA Administrator 

39 Carol M. Browner dated August 22, 1995.
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1

2 List of Subjects in 40 CFR Part 50

3 Environmental protection, Air pollution control, Carbon monoxide, Lead, Nitrogen 

4 dioxide, Ozone, Particulater matter, Sulfur oxides.

5

6 ________________________________

7 Dated: July 12, 2011

8

9

10 Lisa P. Jackson,

11 Administrator

12

13 For the reasons set forth in the preamble, part 50 of chapter 1 of title 40 of the code of 

14 Federal regulations is proposed to be amended as follows:

15

16 PART 50-NATIONAL PRIMARY AND SECONDARY AMBIENT AIR QUALITY 

17 STANDARDS 

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

19 Authority: 42 U.S.C. 7401, et seq.

20
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1 2. Section 50.5 is amended by revising paragraphs (b) and (c) and by adding paragraphs (d) and 

2 (e) to read as follows:

3 §50.5 National secondary ambient air quality standards for sulfur oxides (sulfur dioxide).

4 * * * * *

5 (b) The level of the national secondary 1-hour ambient air quality standard for oxides of sulfur is 

6 75 parts per billion (ppb, which is 1 part in 1,000,000,000), measured in the ambient air as sulfur 

7 dioxide (SO2).

8 (c) The levels of the standards shall be measured by a reference method based on Appendix A-1 

9 or A-2 of this part, or by a Federal Equivalent Method (FEM) designated in accordance with part 

10 53 of this chapter.

11 (d) To demonstrate attainment with the 3-hour secondary standard, the second-highest 3-hour 

12 average must be based upon hourly data that are at least 75 percent complete in each calendar 

13 quarter. A 3-hour block average shall be considered valid only if all three hourly averages for 

14 the 3-hour period are available. If only one or two hourly averages are available, but the 3-hour 

15 average would exceed the level of the standard when zeros are substituted for the missing values, 

16 subject to the rounding rule of paragraph (a) of this section, then this shall be considered a valid 

17 3-hour average. In all cases, the 3-hour block average shall be computed as the sum of the 

18 hourly averages divided by 3.

19 (e) The 1-hour secondary standard is met at an ambient air quality monitoring site when the 

three-year average of the annual 99th 20 percentile of the daily maximum 1-hour average 

21 concentrations is less than or equal to 75 ppb, as determined in accordance with Appendix T of 

22 this part.

23
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1 3. Section 50.11 is amended by deleting paragraphs (f) and (g) and revising paragraphs (a) 

2 through (e) to read as follows: 

3 §50.11 National primary and secondary ambient air quality standards for oxides of 

4 nitrogen (with nitrogen dioxide as the indicator).

5 (a) The level of the national primary and secondary annual ambient air quality standards for 

6 oxides of nitrogen is 53 parts per billion (ppb, which is 1 part in 1,000,000,000), annual average 

7 concentration, measured in the ambient air as nitrogen dioxide.

8 (b) The level of the national primary and secondary 1-hour ambient air quality standards for 

9 oxides of nitrogen is 100 ppb, 1-hour average concentration, measured in the ambient air as 

10 nitrogen dioxide. 

11 (c) The levels of the standards shall be measured by:

12 (1) a reference method based on appendix F to this part; or 

13 (2) a Federal equivalent method (FEM) designated in accordance with part 53 of this chapter.

14 (d) The annual primary and secondary standards are met when the annual average concentration 

15 in a calendar year is less than or equal to 53 ppb, as determined in accordance with Appendix S

16 of this part for the annual standard.

17 (e) The 1-hour primary and secondary standards are met when the three-year average of the 

annual 98th 18 percentile of the daily maximum 1-hour average concentration is less than or equal to 

19 100 ppb, as determined in accordance with Appendix S of this part for the 1-hour standard.

20
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1 4. Appendix S is amended as follows:

2 a. by revising paragraph 1.(a),

3 b. by revising the definition of “Design values” under paragraph 1.(c),

4 c. by revising paragraph 2.(b),

5 d. by revising paragraphs 3.1(a) through (d),

6 e. by revising paragraphs 3.2(a) through (e),

7 f. by revising paragraph 4.1(b),

8 g. by revising paragraph 4.2 (c),

9 h. by revising paragraph 5.1(b), and

10 i. by revising paragraph 5.2(b) to read as follows:

11 Appendix S to Part 50—Interpretation of the Primary and Secondary National Ambient 

12 Air Quality Standards for Oxides of Nitrogen (Nitrogen Dioxide)

13 1. General.

14 (a) This appendix explains the data handling conventions and computations necessary for 

15 determining when the primary and secondary national ambient air quality standards for oxides of 

16 nitrogen as measured by nitrogen dioxide (“NO2 NAAQS”) specified in §50.11 are met. 

17 Nitrogen dioxide (NO2) is measured in the ambient air by a Federal reference method (FRM) 

18 based on appendix F to this part or by a Federal equivalent method (FEM) designated in 

19 accordance with part 53 of this chapter. Data handling and computation procedures to be used in 
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1 making comparisons between reported NO2 concentrations and the levels of the NO2 NAAQS 

2 are specified in the following sections.

3 * * * * * 

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

5 * * * * * 

6 Design values are the metrics (i.e., statistics) that are compared to the NAAQS levels to 

7 determine compliance, calculated as specified in section 5 of this appendix. The design values 

8 for the primary and secondary NAAQS are:

9 (1) The annual mean value for a monitoring site for one year (referred to as the “annual primary

10 or secondary standard design value”). 

(2) The 3-year average of annual 98th 11 percentile daily maximum 1-hour values for a monitoring 

12 site (referred to as the “1-hour primary or secondary standard design value”).

13 * * * * * 

14 2. Requirements for Data Used for Comparisons with the NO2 NAAQS and Data

15 Reporting Considerations.

16 * * * * * 

17 (b) When two or more NO2 monitors are operated at a site, the state may in advance designate 

18 one of them as the primary monitor. If the state has not made this designation, the Administrator 

19 will make the designation, either in advance or retrospectively. Design values will be developed 

20 using only the data from the primary monitor, if this results in a valid design value. If data from 

21 the primary monitor do not allow the development of a valid design value, data solely from the 

22 other monitor(s) will be used in turn to develop a valid design value, if this results in a valid 
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1 design value. If there are three or more monitors, the order for such comparison of the other 

2 monitors will be determined by the Administrator. The Administrator may combine data from 

3 different monitors in different years for the purpose of developing a valid 1-hour primary or 

4 secondary standard design value, if a valid design value cannot be developed solely with the data 

5 from a single monitor. However, data from two or more monitors in the same year at the same 

6 site will not be combined in an attempt to meet data completeness requirements, except if one 

7 monitor has physically replaced another instrument permanently, in which case the two 

8 instruments will be considered to be the same monitor, or if the state has switched the 

9 designation of the primary monitor from one instrument to another during the year.

10 * * * * * 

11 3. Comparisons with the NO2 NAAQS.

12 3.1 The Annual Primary and Secondary NO2 NAAQS.

13 (a) The annual primary and secondary NO2 NAAQS are met at a site when the valid annual 

14 primary standard design value is less than or equal to 53 parts per billion (ppb).

15 (b) An annual primary or secondary standard design value is valid when at least 75 percent of the 

16 hours in the year are reported. 

17 (c) An annual primary or secondary standard design value based on data that do not meet the 

18 completeness criteria stated in section 3.1(b) may also be considered valid with the approval of, 

19 or at the initiative of, the Administrator, who may consider factors such as monitoring site 

20 closures/moves, monitoring diligence, the consistency and levels of the valid concentration 

21 measurements that are available, and nearby concentrations in determining whether to use such 

22 data.
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1 (d) The procedures for calculating the annual primary and secondary standard design values are 

2 given in section 5.1 of this appendix.

3 3.2 The 1-hour Primary and Secondary NO2 NAAQS.

4 (a) The 1-hour primary or secondary NO2 NAAQS is met at a site when the valid 1-hour primary

5 or secondary standard design value is less than or equal to 100 parts per billion (ppb).

6 (b) An NO2 1-hour primary or secondary standard design value is valid if it encompasses three

7 consecutive calendar years of complete data. A year meets data completeness requirements 

8 when all 4 quarters are complete. A quarter is complete when at least 75 percent of the sampling 

9 days for each quarter have complete data. A sampling day has complete data if 75 percent of the 

10 hourly concentration values, including state-flagged data affected by exceptional events which 

11 have been approved for exclusion by the Administrator, are reported.

12 (c) In the case of one, two, or three years that do not meet the completeness requirements of 

13 section 3.2(b) of this appendix and thus would normally not be useable for the calculation of a 

14 valid 3-year 1-hour primary or secondary standard design value, the 3-year 1-hour primary or 

15 secondary standard design value shall nevertheless be considered valid if one of the following 

16 conditions is true. 

17 (i) At least 75 percent of the days in each quarter of each of three consecutive years have at least 

18 one reported hourly value, and the design value calculated according to the procedures specified 

19 in section 5.2 is above the level of the primary or secondary 1-hour standard.

20 (ii) (A) A 1-hour primary or secondary standard design value that is below the level of the 

21 NAAQS can be validated if the substitution test in section 3.2(c)(ii)(B) results in a “test design 

22 value” that is below the level of the NAAQS. The test substitutes actual ‘‘high’’ reported daily

23 maximum 1-hour values from the same site at about the same time of the year (specifically, in 
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1 the same calendar quarter) for unknown values that were not successfully measured. Note that 

2 the test is merely diagnostic in nature, intended to confirm that there is a very high likelihood 

3 that the original design value (the one with less than 75 percent data capture of hours by day and 

4 of days by quarter) reflects the true under-NAAQS-level status for that 3-year period; the result 

5 of this data substitution test (the ‘‘test design value”, as defined in section 3.2(c)(ii)(B)) is not 

6 considered the actual design value. For this test, substitution is permitted only if there are at least 

7 200 days across the three matching quarters of the three years under consideration (which is 

8 about 75 percent of all possible daily values in those three quarters) for which 75 percent of the 

9 hours in the day, including state-flagged data affected by exceptional events which have been 

10 approved for exclusion by the Administrator, have reported concentrations. However, maximum 

11 1-hour values from days with less than 75 percent of the hours reported shall also be considered 

12 in identifying the high value to be used for substitution.

13 (B) The substitution test is as follows: Data substitution will be performed in all quarter periods 

14 that have less than 75 percent data capture but at least 50 percent data capture, including state15 flagged data affected by exceptional events which have been approved for exclusion by the 

16 Administrator; if any quarter has less than 50 percent data capture then this substitution test

17 cannot be used. Identify for each quarter (e.g., January-March) the highest reported daily

18 maximum 1-hour value for that quarter, excluding state-flagged data affected by exceptional 

19 events which have been approved for exclusion by the Administrator, looking across those three 

20 months of all three years under consideration. All daily maximum 1-hour values from all days in 

21 the quarter period shall be considered when identifying this highest value, including days with 

22 less than 75 percent data capture. If after substituting the highest non-excluded reported daily 

23 maximum 1-hour value for a quarter for as much of the missing daily data in the matching 
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1 deficient quarter(s) as is needed to make them 100 percent complete, the procedure in section 5.2

2 yields a recalculated 3-year 1-hour standard “test design value” below the level of the standard, 

3 then the 1-hour primary or secondary standard design value is deemed to have passed the 

4 diagnostic test and is valid, and the level of the standard is deemed to have been met in that 3-

5 year period. As noted in section 3.2(c)(i), in such a case, the 3-year design value based on the 

6 data actually reported, not the ‘‘test design value’’, shall be used as the valid design value.

7 (iii) (A) A 1-hour primary or secondary standard design value that is above the level of the 

8 NAAQS can be validated if the substitution test in section 3.2(c)(iii)(B) results in a “test design 

9 value” that is above the level of the NAAQS. The test substitutes actual ‘‘low” reported daily 

10 maximum 1-hour values from the same site at about the same time of the year (specifically, in 

11 the same three months of the calendar) for unknown values that were not successfully measured. 

12 Note that the test is merely diagnostic in nature, intended to confirm that there is a very high 

13 likelihood that the original design value (the one with less than 75 percent data capture of hours 

14 by day and of days by quarter) reflects the true above-NAAQS-level status for that 3-year period; 

15 the result of this data substitution test (the ‘‘test design value”, as defined in section 

16 3.2(c)(iii)(B)) is not considered the actual design value. For this test, substitution is permitted 

17 only if there are a minimum number of available daily data points from which to identify the low

18 quarter-specific daily maximum 1-hour values, specifically if there are at least 200 days across 

19 the three matching quarters of the three years under consideration (which is about 75 percent of

20 all possible daily values in those three quarters) for which 75 percent of the hours in the day have 

21 reported concentrations. Only days with at least 75 percent of the hours reported shall be 

22 considered in identifying the low value to be used for substitution.
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1 (B) The substitution test is as follows: Data substitution will be performed in all quarter periods 

2 that have less than 75 percent data capture. Identify for each quarter (e.g., January-March) the 

3 lowest reported daily maximum 1-hour value for that quarter, looking across those three months 

4 of all three years under consideration. All daily maximum 1-hour values from all days with at 

5 least 75 percent capture in the quarter period shall be considered when identifying this lowest 

6 value. If after substituting the lowest reported daily maximum 1-hour value for a quarter for as 

7 much of the missing daily data in the matching deficient quarter(s) as is needed to make them 75

8 percent complete, the procedure in section 5.2 yields a recalculated 3-year 1-hour standard “test 

9 design value” above the level of the standard, then the 1-hour primary or secondary standard 

10 design value is deemed to have passed the diagnostic test and is valid, and the level of the 

11 standard is deemed to have been exceeded in that 3-year period. As noted in section 3.2(c)(i), in 

12 such a case, the 3-year design value based on the data actually reported, not the ‘‘test design 

13 value’’, shall be used as the valid design value.

14 (d) A 1-hour primary or secondary standard design value based on data that do not meet the 

15 completeness criteria stated in 3.2(b) and also do not satisfy section 3.2(c), may also be 

16 considered valid with the approval of, or at the initiative of, the Administrator, who may consider 

17 factors such as monitoring site closures/moves, monitoring diligence, the consistency and levels 

18 of the valid concentration measurements that are available, and nearby concentrations in 

19 determining whether to use such data.

20 (e) The procedures for calculating the 1-hour primary and secondary standard design values are 

21 given in section 5.2 of this appendix.

22 4. Rounding Conventions. 

23 4.1 Rounding Conventions for the Annual Primary and Secondary NO2 NAAQS.
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1 * * * * *

2 (b) The annual primary or secondary standard design value is calculated pursuant to section 5.1 

3 and then rounded to the nearest whole number or 1 ppb (decimals 0.5 and greater are rounded up 

4 to the nearest whole number, and any decimal lower than 0.5 is rounded down to the nearest 

5 whole number).

6 4.2 Rounding Conventions for the 1-hour Primary and Secondary NO2 NAAQS.

7 * * * * *

8 (c) The 1-hour primary or secondary standard design value is calculated pursuant to section 5.2 

9 and then rounded to the nearest whole number or 1 ppb (decimals 0.5 and greater are rounded up 

10 to the nearest whole number, and any decimal lower than 0.5 is rounded down to the nearest 

11 whole number).

12 5. Calculation Procedures for the Primary and Secondary NO2 NAAQS.

13 5.1 Procedures for the Annual Primary and Secondary NO2 NAAQS

14 * * * * *

15 (b) The annual primary or secondary standard design value for a site is the valid annual mean 

16 rounded according to the conventions in section 4.1.

17 5.2 Calculation Procedures for the 1-hour Primary and Secondary NO2 NAAQS.

18 * * * * *

19 (b) The 1-hour primary or secondary standard design value for a site is the mean of the three 

annual 98th 20 percentile values, rounded according to the conventions in section 4.

21 * * * * *

22
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1 5. Appendix T is amended as follows:

2 a. by revising paragraph 1.(a),

3 b. by revising the definition of “Design values” under paragraph 1.(c),

4 c. by revising paragraph 2.(b),

5 d. by revising paragraphs 3.(a) through (e),

6 e. by revising paragraph 4.(c), and

7 f. by revising paragraph 5.(b) to read as follows:

8 Appendix T to Part 50—Interpretation of the Primary and Secondary National Ambient 

9 Air Quality Standards for Oxides of Sulfur (Sulfur Dioxide) 

10 1. General.

11 (a) This appendix explains the data handling conventions and computations necessary for 

12 determining when the primary and secondary national ambient air quality standards for Oxides 

13 of Sulfur as measured by Sulfur Dioxide (“SO2 NAAQS”) specified in § 50.17 and §50.5 (b), 

14 respectively, are met at an ambient air quality monitoring site. Sulfur Dioxide (SO2) is measured 

15 in the ambient air by a Federal reference method (FRM) based on appendix A-1 or A-2 to this 

16 part or by a Federal equivalent method (FEM) designated in accordance with part 53 of this 

17 chapter. Data handling and computation procedures to be used in making comparisons between 

18 reported SO2 concentrations and the levels of the SO2 NAAQS are specified in the following 

19 sections.
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1 * * * * *

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

3 * * * * *

4 Design values are the metrics (i.e., statistics) that are compared to the NAAQS levels to 

5 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 6 percentile daily 

7 maximum 1-hour values for a monitoring site (referred to as the “1-hour primary standard design 

8 value”).

9 * * * * *

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

11 Reporting Considerations.

12 * * * * *

13 (b) Data from two or more monitors from the same year at the same site reported to EPA under 

14 distinct Pollutant Occurrence Codes shall not be combined in an attempt to meet data 

completeness requirements. The Administrator will combine annual 99th 15 percentile daily 

16 maximum concentration values from different monitors in different years, selected as described 

17 here, for the purpose of developing a valid 1-hour primary or secondary standard design value. 

18 If more than one of the monitors meets the completeness requirement for all four quarters of a 

19 year, the steps specified in section 5(a) of this appendix shall be applied to the data from the 

20 monitor with the highest average of the four quarterly completeness values to derive a valid 

annual 99th 21 percentile daily maximum concentration. If no monitor is complete for all four 

22 quarters in a year, the steps specified in section 3(c) and 5(a) of this appendix shall be applied to 
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1 the data from the monitor with the highest average of the four quarterly completeness values in 

an attempt to derive a valid annual 99th 2 percentile daily maximum concentration. This paragraph 

3 does not prohibit a monitoring agency from making a local designation of one physical monitor 

4 as the primary monitor for a Pollutant Occurrence Code and substituting the 1-hour data from a 

5 second physical monitor whenever a valid concentration value is not obtained from the primary

6 monitor; if a monitoring agency substitutes data in this manner, each substituted value must be 

7 accompanied by an AQS qualifier code indicating that substitution with a value from a second 

8 physical monitor has taken place.

9 * * * * *

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

11 (a) The 1-hour primary or secondary SO2 NAAQS is met at an ambient air quality monitoring 

12 site when the valid 1-hour primary or secondary standard design value is less than or equal to 75

13 parts per billion (ppb).

14 (b) An SO2 1-hour primary or secondary standard design value is valid if it encompasses three

15 consecutive calendar years of complete data. A year meets data completeness requirements 

16 when all 4 quarters are complete. A quarter is complete when at least 75 percent of the sampling 

17 days for each quarter have complete data. A sampling day has complete data if 75 percent of the 

18 hourly concentration values, including State-flagged data affected by exceptional events which 

19 have been approved for exclusion by the Administrator, are reported.

20 (c) In the case of one, two, or three years that do not meet the completeness requirements of 

21 section 3(b) of this appendix and thus would normally not be useable for the calculation of a 

22 valid 3-year 1-hour primary or secondary standard design value, the 3-year 1-hour primary or 
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1 secondary standard design value shall nevertheless be considered valid if one of the following 

2 conditions is true. 

3 (i) At least 75 percent of the days in each quarter of each of three consecutive years have at least 

4 one reported hourly value, and the design value calculated according to the procedures specified 

5 in section 5 is above the level of the primary or secondary 1-hour standard.

6 (ii) (A) A 1-hour primary or secondary standard design value that is equal to or below the level 

7 of the NAAQS can be validated if the substitution test in section 3(c)(ii)(B) results in a “test 

8 design value” that is below the level of the NAAQS. The test substitutes actual ‘‘high’’ reported 

9 daily maximum 1-hour values from the same site at about the same time of the year (specifically, 

10 in the same calendar quarter) for unknown values that were not successfully measured. Note that 

11 the test is merely diagnostic in nature, intended to confirm that there is a very high likelihood 

12 that the original design value (the one with less than 75 percent data capture of hours by day and 

13 of days by quarter) reflects the true under-NAAQS-level status for that 3-year period; the result 

14 of this data substitution test (the ‘‘test design value”, as defined in section 3(c)(ii)(B)) is not 

15 considered the actual design value. For this test, substitution is permitted only if there are at least 

16 200 days across the three matching quarters of the three years under consideration (which is 

17 about 75 percent of all possible daily values in those three quarters) for which 75 percent of the 

18 hours in the day, including State-flagged data affected by exceptional events which have been 

19 approved for exclusion by the Administrator, have reported concentrations. However, maximum 

20 1-hour values from days with less than 75 percent of the hours reported shall also be considered 

21 in identifying the high value to be used for substitution.

22 (B) The substitution test is as follows: Data substitution will be performed in all quarter periods 

23 that have less than 75 percent data capture but at least 50 percent data capture, including State-
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1 flagged data affected by exceptional events which have been approved for exclusion by the 

2 Administrator; if any quarter has less than 50 percent data capture then this substitution test

3 cannot be used. Identify for each quarter (e.g., January-March) the highest reported daily

4 maximum 1-hour value for that quarter, excluding State-flagged data affected by exceptional 

5 events which have been approved for exclusion by the Administrator, looking across those three 

6 months of all three years under consideration. All daily maximum 1-hour values from all days in 

7 the quarter period shall be considered when identifying this highest value, including days with 

8 less than 75 percent data capture. If after substituting the highest reported daily maximum 1-hour 

9 value for a quarter for as much of the missing daily data in the matching deficient quarter(s) as is 

10 needed to make them 100 percent complete, the procedure in section 5 yields a recalculated 3-

11 year 1-hour standard “test design value” less than or equal to the level of the standard, then the 1-

12 hour primary or secondary standard design value is deemed to have passed the diagnostic test 

13 and is valid, and the level of the standard is deemed to have been met in that 3-year period. As 

14 noted in section 3(c)(i), in such a case, the 3-year design value based on the data actually

15 reported, not the ‘‘test design value’’, shall be used as the valid design value.

16 (iii) (A) A 1-hour primary or secondary standard design value that is above the level of the 

17 NAAQS can be validated if the substitution test in section 3(c)(iii)(B) results in a “test design 

18 value” that is above the level of the NAAQS. The test substitutes actual ‘‘low” reported daily 

19 maximum 1-hour values from the same site at about the same time of the year (specifically, in 

20 the same three months of the calendar) for unknown hourly values that were not successfully

21 measured. Note that the test is merely diagnostic in nature, intended to confirm that there is a 

22 very high likelihood that the original design value (the one with less than 75 percent data capture

23 of hours by day and of days by quarter) reflects the true above-NAAQS-level status for that 3-
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1 year period; the result of this data substitution test (the ‘‘test design value”, as defined in section 

2 3(c)(iii)(B)) is not considered the actual design value. For this test, substitution is permitted only

3 if there are a minimum number of available daily data points from which to identify the low

4 quarter-specific daily maximum 1-hour values, specifically if there are at least 200 days across 

5 the three matching quarters of the three years under consideration (which is about 75 percent of 

6 all possible daily values in those three quarters) for which 75 percent of the hours in the day have 

7 reported concentrations. Only days with at least 75 percent of the hours reported shall be 

8 considered in identifying the low value to be used for substitution.

9 (B) The substitution test is as follows: Data substitution will be performed in all quarter periods 

10 that have less than 75 percent data capture. Identify for each quarter (e.g., January-March) the 

11 lowest reported daily maximum 1-hour value for that quarter, looking across those three months 

12 of all three years under consideration. All daily maximum 1-hour values from all days with at 

13 least 75 percent capture in the quarter period shall be considered when identifying this lowest 

14 value. If after substituting the lowest reported daily maximum 1-hour value for a quarter for as 

15 much of the missing daily data in the matching deficient quarter(s) as is needed to make them 75

16 percent complete, the procedure in section 5 yields a recalculated 3-year 1-hour standard “test 

17 design value” above the level of the standard, then the 1-hour primary or secondary standard 

18 design value is deemed to have passed the diagnostic test and is valid, and the level of the 

19 standard is deemed to have been exceeded in that 3-year period. As noted in section 3(c)(i), in 

20 such a case, the 3-year design value based on the data actually reported, not the ‘‘test design 

21 value’’, shall be used as the valid design value.

22 (d) A 1-hour primary or secondary standard design value based on data that do not meet the 

23 completeness criteria stated in 3(b) and also do not satisfy section 3(c), may also be considered 
Interagency Working Comments on Draft Language under EO12866 and 

13563 Interagency Review. Subject to Further Policy Review.

238

1 valid with the approval of, or at the initiative of, the Administrator, who may consider factors 

2 such as monitoring site closures/moves, monitoring diligence, the consistency and levels of the 

3 valid concentration measurements that are available, and nearby concentrations in determining 

4 whether to use such data.

5 (e) The procedures for calculating the 1-hour primary or secondary standard design values are 

6 given in section 5 of this appendix.

7 4. Rounding Conventions for the 1-hour Primary and Secondary SO2 NAAQS. 

8 * * * * * 

9 (c) The 1-hour primary or secondary standard design value is calculated pursuant to section 5 

10 and then rounded to the nearest whole number or 1 ppb (decimals 0.5 and greater are rounded up 

11 to the nearest whole number, and any decimal lower than 0.5 is rounded down to the nearest 

12 whole number).

13 5. Calculation Procedures for the 1-hour Primary and Secondary SO2 NAAQS.

14 * * * * *

15 (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 16 percentile values, rounded according to the conventions 

17 in section 4.