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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 set a new multi17 pollutant secondary standard in this review. The new 1-hour secondary NO2 standard is set at a 

18 level of 100 ppb and the new 1-hour secondary SO2 standard is set at 75 ppb. In addition, EPA 

19 has decided to undertake a field pilot program to gather and analyze additional relevant data so 

20 as to enhance the Agency’s understanding of the degree of protectiveness that a new multi21 pollutant approach, defined in terms of an aquatic acidification index (AAI), would afford and to 

22 support development of an appropriate monitoring network for such a standard. The EPA 
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1 solicits comment on the framework of such a standard and on the design of the field pilot 

2 program.

3 DATES: Written comments on this proposed rule must be received by [insert date 60 days 

4 after date of publication in the Federal Register].

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

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

7 address of the public hearing on this proposed rule.

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

9 1145, by one of the following methods:

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

11  Email: a-and-r-Docket@epa.gov.

12  Fax: 202-566-1741.

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

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

15 total of two copies.

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

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

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

19 arrangements should be made for deliveries of boxed information.

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

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

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

23 provided, unless the comment includes information claimed to be Confidential Business 
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1 Information (CBI) or other information whose disclosure is restricted by statute. Do not submit 

2 information that you consider to be CBI or otherwise protected through www.regulations.gov or 

3 email. The www.regulations.gov website is an “anonymous access” system, which means EPA 

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

22 through Friday, excluding legal holidays. The telephone number for the Public Reading Room is 
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1 (202) 566-1744 and the telephone number for the Air and Radiation Docket and Information 

2 Center is (202) 566-1742. 

3 FOR FURTHER INFORMATION CONTACT: Dr. Richard Scheffe, Office of Air Quality

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

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

6 scheffe.rich@epa.gov. 

7 SUPPLEMENTARY INFORMATION:

8 General Information

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

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

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

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

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

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

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

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

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

18 forth in 40 CFR part 2.

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

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

21 heading, Federal Register date and page number).
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1  Follow directions – The Agency may ask you to respond to specific questions or 

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

3 number.

4  Explain why you agree or disagree, suggest alternatives, and substitute language for your 

5 requested changes.

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

7 used.

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

9 sufficient detail to allow for it to be reproduced.

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

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

12 comment period deadline identified.

13 Availability of Related Information

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

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

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

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

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

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

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

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

22 risk assessments and other related technical documents are available at 
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1 http://www.epa.gov/ttn/naaqs/standards/no2so2sec/cr_rea.html. These and other related 

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

3 Table of Contents

4 The following topics are discussed in this preamble:

5 I. Background

6 A. Legislative Requirements

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

8 1. NAAQS for Oxides of Nitrogen

9 2. NAAQS for Oxides of Sulfur

10 C. History of Related Assessments and Agency Actions

11 D. History of the Current Review

12 E. Scope of the Current Review

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

14 A. Ecological Effects

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

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

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

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

19 and Sulfur

20 a. Nature of Acidification –related Ecosystem Responses

21 i. Aquatic ecosystems

22 ii. Terrestrial ecosystems

23 iii. Ecosystem sensitivity

24 b. Magnitude of Acidification-related Ecosystem Responses

25 i. Aquatic acidification

26 ii. Terrestrial acidification

27 c. Key Uncertainties Associated with Acidification

28 i. Aquatic acidification

29 ii. Terrestrial acidification

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

31 Nitrogen 

32 a. Nature of Nutrient Enrichment-related Ecosystem Responses

33 i. Aquatic ecosystems

34 ii. Terrestrial ecosystems

35 iii. Ecosystem sensitivity to nutrient enrichment

36 b. Magnitude of Nutrient Enrichment-related Ecosystem Responses

37 i. Aquatic ecosystems

38 ii. Terrestrial ecosystems

39 c. Key Uncertainties Associated with Nutrient Enrichment

40 i. Aquatic ecosystems

41 ii. Terrestrial ecosystems

42 4. Other Ecological Effects
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1 B. Risk and Exposure Assessment

2 1. Overview of Risk and Exposure Assessment

3 2. Key Findings

4 a. Air quality analyses

5 b. Deposition-related aquatic acidification

6 c. Deposition-related terrestrial acidification

7 d. Deposition-related aquatic nutrient enrichment

8 e. Deposition-related terrestrial nutrient enrichment 

9 f. Additional effects

10 3. Conclusions on Effects

11 C. Adversity of Effects to Public Welfare

12 1. Ecosystem Services

13 2. Effects on Ecosystem Services

14 a. Aquatic Acidification

15 b. Terrestrial Acidification

16 c. Nutrient Enrichment

17 3. Summary

18 D. Adequacy of the Current Standards

19 1. Adequacy of the Current Standards for Direct Effects

20 2. Appropriateness and Adequacy of the Current Standards for Deposition21 related Effects

22 a. Appropriateness

23 b. Adequacy of Protection

24 i. Aquatic Acidification

25 ii. Terrestrial acidification

26 iii. Terrestrial nutrient enrichment

27 iv. Aquatic nutrient enrichment

28 v. Other effects

29 3. CASAC Views

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

31 Standard

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

33 Standards for Aquatic Acidification

34 A. Ambient Air Indicators

35 1. Oxides of Sulfur

36 2. Oxides of Nitrogen

37 B. Form

38 1. Ecological Indicator

39 2. Linking ANC to Deposition

40 3. Linking Deposition to Ambient Air Indicators

41 4. Aquatic Acidification Index

42 5. Spatial Aggregation

43 a. Ecoregion Sensitivity

44 b. Representative Ecoregion-specific Factors

45 i. Factor F1

46 a) Acid-sensitive Ecoregions
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1 b) Non-acid sensitive Ecoregions

2 ii. Factor F2

3 iii. Factors F3 and F4

4 c. Factors in Data-limited Ecoregions

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

6 6. Summary of the AAI Form

7 C. Averaging Time

8 D. Level

9 1. Association Between pH Levels and Target ANC Levels 

10 2. ANC Levels Related to Effects on Aquatic Ecosystems 

11 3. Consideration of Episodic Acidity

12 4. Consideration of Ecosystem Response Time 

13 5. Prior Examples of Target ANC Levels 

14 6. Consideration of Public Welfare Benefits

15 7. Summary of Alternative Levels

16 E. Combined Alternative Levels and Forms

17 F. Characterization of Uncertainties

18 1. Overview of Uncertainty

19 2. Uncertainties Associated with Data Gaps

20 3. Uncertainties in Modeled Processes

21 G. CASAC Advice

22 H. Administrator’s Proposed Conclusions

23 IV. Field Pilot Program and Ambient Monitoring 

24 A. Field Pilot Program

25 a. Objectives

26 b. Overview of Field Pilot Program

27 c. Complementary Measurements

28 d. Complementary Research Efforts

29 e. Implementation Challenges

30 f. Final Monitoring Plan Development and Stakeholder Participation

31 B. Evaluation of Monitoring Methods

32 1. Potential FRMs for SO2 and p-SO4

33 2. Potential FRM for NOy

34 V. Statutory and Executive Order Reviews

35 A. Executive Order 12866: Regulatory Planning and Review

36 B. Paperwork Reduction Act

37 C. Regulatory Flexibility Act

38 D. Unfunded Mandates Reform Act

39 E. Executive Order 13132: Federalism

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

41 Governments

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

43 Safety Risks

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

45 Distribution, or Use

46 I. National Technology Transfer and Advancement Act
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1 J. Executive Order 12898: Federal Actions to Address Environmental Justice in 

2 Minority Populations and Low-Income Populations

3 References

4

5 I. Background

6 A. Legislative Requirements

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

 

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

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

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

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

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

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

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

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

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

10 472, 475-76 (2001). Likewise, “[a]ttainability and technological feasibility are not relevant 

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

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

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

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

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

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

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

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

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

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

21 Scientific Advisory Committee (CASAC).

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

23 1. NAAQS for Oxides of Nitrogen
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1 After reviewing the relevant science on the public health and welfare effects associated 

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

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

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

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

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

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

8 25532).

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

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

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

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

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

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

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

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

17 make a decision as to the appropriate NAAQS for NO2” (Wolff, 1993). The Air Quality Criteria 

18 for Oxides of Nitrogen was then finalized (US EPA, 1995a). EPA’s OAQPS also prepared a 

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

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

21 the review of the NO2 NAAQS. CASAC reviewed two drafts of the Staff Paper and concluded 

22 in a closure letter to the Administrator that the document provided a “scientifically adequate 

23 basis for regulatory decisions on nitrogen dioxide” (Wolff, 1995).
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1 In October 1995 the Administrator announced her proposed decision not to revise either 

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

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

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

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

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

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

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

9 concentrations.

10 2. NAAQS for Oxides of Sulfur

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

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

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

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

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

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

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

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

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

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

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

22 visibility, soils, and water. However, the available data were considered insufficient to establish 
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1 a quantitative relationship between specific ambient concentrations of oxides of sulfur and such 

2 public welfare effects (38 FR 25679).

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

23 Effects: Critical Assessment Document (US EPA, 1985) (53 FR 14935 -14936). These 
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1 documents, though they were not considered criteria documents and did not undergo CASAC 

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

3 deposition completed by EPA at that point.

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

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

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

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

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

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

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

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

12 annual standards (75 FR 35520), no further decisions on the secondary SO2 standard have been 

13 published. 

14 C. History of Related Assessments and Agency Actions

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

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

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

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

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

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

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

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

23 needed. The Clean Air Act Amendments of 1990 included numerous separate provisions related 
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1 to the acid deposition problem. The primary and most important of the provisions, the 

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

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

4 levels in order to achieve reductions over broad geographic regions. In this provision, Congress 

5 included a statement of findings that led them to take action, concluding that (1) the presence of 

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

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

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

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

10 problem. 

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

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

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

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

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

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

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

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

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

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

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

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

23 standards for sulfur and nitrogen deposition may at some point in the future be technically
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1 feasible, although appropriate deposition loads for these acidifying chemicals could not be 

2 defined with reasonable certainty at that time. 

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

4 pertaining to the scope and application of the secondary NAAQS designed to protect the public 

5 welfare. Specifically, the definition of “effects on welfare” in Section 302(h) was expanded to 

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

7 combination with other air pollutants.” 

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

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

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

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

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

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

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

15 109 of the Act.” The petition further stated that “recent federal studies, including the NAPAP 

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

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

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

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

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

21 warming, and tropospheric ozone and stratospheric ozone depletion.

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

23 requested in 2000 that EPA initiate a rulemaking proceeding to enhance the air quality in 
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1 national parks and wilderness areas in order to protect resources and values that are being 

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

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

4 waters; visibility impairment; and foliar injury from ozone.

5 In a Federal Register notice in 2001, EPA announced receipt of these requests and asked 

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

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

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

9 48699).

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

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

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

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

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

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

16 2005).

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

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

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

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

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

22 pollutants.

23 D. History of the Current Review
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1 The EPA initiated this current review in December 2005 with a call for information (70 

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

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

4 Review Plan (IRP) was developed to provide the framework and schedule as well as the scope of 

5 the review and to identify policy-relevant questions to be addressed in the components of the 

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

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

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

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

10 2-3, 2008 to review the first draft ISA. A second draft ISA was released for CASAC and public 

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

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

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

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

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

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

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

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

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

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

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

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

23 advice and recommendations to the Agency concerning the first draft Policy Assessment (Russell 
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1 and Samet, 2010a). A second draft Policy Assessment was released to CASAC and the public 

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

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

4 December 9, 2010 letter (Russell and Samet 2010b). CASAC and public comments on the 

5 second draft Policy Assessment were considered by EPA staff in developing a final Policy

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

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

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

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

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

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

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

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

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

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

by statute.216 The schedule for completion of this review is governed by a consent decree 

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

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

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

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

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

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

 

2 Center for Biological Diversity, et al. v. Johnson, No. 05–1814 (D.D.C.)
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1 This action presents the Administrator’s proposed decisions on the review of the current 

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

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

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

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

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

7 conclusions, findings, and determinations.

8 E. Scope of the Current Review

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

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

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

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

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

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

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

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

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

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

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

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

21 these interactions and associated ecological effects.

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

23 standards for two other criteria pollutants related to oxides of nitrogen and sulfur: ozone and 
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1 particulate matter (PM). Oxides of nitrogen are precursors to the formation of ozone in the 

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

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

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

5 are a number of welfare effects directly associated with ozone and fine PM, including ozone6 related damage to vegetation and PM-related visibility impairment. Protection against those 

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

23 differently from a public welfare perspective than are the effects of nitrogen deposition in 
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1 natural, unmanaged ecosystems, largely due to the more homogeneous, controlled nature of 

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

3 increased productivity in those ecosystems.

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

5 the effects of ambient oxides of nitrogen and sulfur via deposition on multiple ecological 

6 receptors. The ISA highlights effects including those associated with acidification and nitrogen 

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

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

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

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

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

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

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

14 with the presence of these air pollutants in the ambient air. Consistent with the CAA, this policy

15 objective includes consideration of “variable factors . . . which of themselves or in combination 

16 with other factors may alter the effects on public welfare” of the criteria air pollutants included in 

17 this review.

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

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

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

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

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

23 2010b), and as discussed in detail in the Policy Assessment, we have the greatest confidence in 
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1 the causal linkages between oxides of nitrogen and sulfur and aquatic acidification effects 

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

3 terrestrial nutrient enrichment.

4

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

6 Standards

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

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

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

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

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

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

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

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

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

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

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

18 for this purpose.

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

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

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

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

23 nitrogen and oxides of sulfur in the ambient air. This rationale also takes into account: (1) staff 
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1 assessments of the most policy-relevant information in the ISA and staff analyses of air quality, 

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

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

4 sulfur standards are based; (2) CASAC advice and recommendations, as reflected in discussions 

5 of drafts of the ISA, REA, and Policy Assessment at public meetings, in separate written 

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

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

8 separately. 

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

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

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

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

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

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

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

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

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

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

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

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

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

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

23 amount of deposition that a given ecosystem can accept without defined levels of degradation.
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1 Drawing from the framework developed in the REA, the framework we used to structure 

2 an ecologically meaningful secondary standard in the 

3 the indicator, form, level, and averaging time of such as standard in section III of this proposal is 

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

5 ecologically relevant standard.

6

7

8 Figure II-1. Simplified conceptual design of

9 for oxides of nitrogen and sulfur.

10

11 The following discussion

12 The Policy Assessment includes staff’s evaluation of the policy implications of the sc

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

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

15 information informs staff conclusions and the development of

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

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

18 2 of the Policy Assessment presents

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

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

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

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

el, and averaging time of such as standard in section III of this proposal is 

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

Simplified conceptual design of the form of an aquatic acidification standard 

for oxides of nitrogen and sulfur.

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

includes staff’s evaluation of the policy implications of the sc

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

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

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

for consideration in addressing public and welfare effects associated with the 

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

presents information not repeated here that characterizes emissions, 

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

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

Interagency Working Comments on Draft Language under EO12866 and 

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

and to further develop 

el, and averaging time of such as standard in section III of this proposal is 

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

the form of an aquatic acidification standard 

Policy Assessment. 

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

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

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

policy options in the Policy

for consideration in addressing public and welfare effects associated with the 

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

information not repeated here that characterizes emissions, 

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

sulfur in the atmosphere as well as current ambient air quality monitoring networks and models. 
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1 Additional information in this section includes ecological modeling and water quality data 

2 sources. 

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

4 sulfur in the ambient air. The discussion is organized around the types of effects being 

5 considered, including direct effects of gaseous oxides of nitrogen and sulfur, deposition-related 

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

7 damage, climate-related effects and mercury methylation.

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

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

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

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

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

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

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

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

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

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

18 linked to ecological indicators.

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

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

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

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

23 providing protection from deposition-related effects. Considerations as to the extent to which 
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1 deposition-related effects that could reasonably be judged to be adverse to public welfare are 

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

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

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

5 A. Ecological Effects

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

23 decreased growth. The following summary is a concise overview of the known or anticipated 
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1 effects to vegetation caused by gas phase nitrogen and sulfur. Most phototoxic effects associated 

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

17 sensitive lichen species. 

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

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

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

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

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

23 3.4.2). The functional relationship between ambient concentrations of gas phase oxides of 
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1 nitrogen and sulfur and specific plant response are impacted by internal factors such as rate of 

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

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

4 3.4.2).

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

23 and HNO3 at concentrations currently observed in the United States with the exception of some 
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1 lichen species (US EPA, 2008, section 4.4). Past and current HNO3 concentrations may be 

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

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

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

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

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

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

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

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

10 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-

11

) and nitrate (NO3

-

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

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

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

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

16 species, including NHx.

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

2-

or NO3

-

17 , 

18 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+ 19 ), 

magnesium (Mg2+), sodium (Na+), or potassium (K+), rather than hydrogen (H+20 ) and dissolved 

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

decreases. Continued SO4

2-

or NO3

-

22 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

-

23 (with 
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H

+

and Al3+ 1 ) leads to acidification of soil water, and by connection, surface water. Introduction 

2 of strong acid anions such as sulfate and nitrate to an already acidic soil, whether naturally or due 

3 to anthropogenic activities, can lead to instantaneous acidification of waterbodies through direct 

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

19 2009, section 4.2). In addition, the accumulated sulfate and nitrate in snow packs can provide a 

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

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

22 consequences to sensitive biota.
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1 The following summary is a concise overview of the known or anticipated effects caused 

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

3 freshwater aquatic ecosystems. 

4 a. Nature of acidification-related ecosystem responses

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

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

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

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

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

10 i. Aquatic ecosystems

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

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

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

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

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

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

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

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

19 between ANC and the surface water constituents that directly contribute to or ameliorate acidityrelated stress, in particular, concentrations of hydrogen ion (as pH), Ca2+ 20 , and aluminum. 

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

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

23 and phyto-zooplankton diversity and abundance. 
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1 Low ANC coincides with effects on aquatic systems (e.g., individual species fitness loss

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

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

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

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

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

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

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

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

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

11 zooplankton) also begins to decline. When ANC concentrations are <50 μeq/L, they are 

12 generally associated with death or loss of fitness of biota that are sensitive to acidification.

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

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

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

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

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

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

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

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

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

22 of acidification (US EPA, 2011, section 3.1.3).
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1 There are clear associations between ANC, pH, and aquatic species mortality and health 

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

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

4 no toxicity range from 6 to 7 (MacAvoy et al, 1995). Baker et al (1990) observed that “lakes 

5 with pH less than approximately 6.0 contain significantly fewer species than lakes with pH levels 

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

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

8 populations, whereas other taxa are reduced to only acidophilic species. Acid Neutralizing 

9 Capacity is a measure of how much acid can be neutralized in a specific surface water system. 

10 An ANC value of 0 or below means that surface waters have no ability to neutralize any

11 additional acid inputs.

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

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

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

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

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

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

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

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

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

21 2008, section 3.2.3.3). 

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

23 pH is not uniform across species. A number of synoptic surveys indicated loss of species 
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1 diversity and absence of several fish species in the pH range of 5.0 to 5.5. If pH is lower, there is 

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

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

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

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

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

7 pH levels are at or below 5.5. 

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

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

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

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

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

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

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

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

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

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

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

19 section 3.2.3.3).

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

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

22 53 potential species). There is often a positive relationship between pH and number of fish 

23 species, at least for pH values between about 5.0 and 6.5, or ANC values between about 0 to 
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1 100 μeq/L. Such observed relationships are complicated, however, by the tendency for smaller 

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

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

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

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

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

7 of zooplankton, and macroinvertebrates which are often sources of food for fish, birds and other 

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

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

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

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

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

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

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

15 Adirondacks. 

16 ii. Terrestrial ecosystems

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

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

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

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

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

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

23 soil water and soil C:N ratio (US EPA, 2008, section 3.2.2.2). 
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1 As discussed in the ISA and REA, in soils with base saturation less than about 15 to 20%, 

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

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

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

The Ca2+ 5 and Al concentrations in soil water are strongly influenced by soil acidification and 

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

Ca2+ 7 uptake and Al toxicity to roots. Effects of nitrification and associated acidification and 

8 cation leaching have been consistently shown to occur only in soils with a C:N ratio below about 

9 20 to 25.

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

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

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

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

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

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

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

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

18 generally rich in base cations.

19 iii. Ecosystem sensitivity 

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

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

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

23 acidification. The ISA reports that the principal factor governing the sensitivity of terrestrial and 
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1 aquatic ecosystems to acidification from sulfur and nitrogen deposition is geology (particularly

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

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

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

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

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

7 b. Magnitude of acidification-related ecosystem responses

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

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

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

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

12 i. Aquatic acidification

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

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

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

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

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

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

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

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

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

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

23 exception of the Los Angeles Basin and surrounding areas, the levels of acidifying deposition are 
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1 low in most areas. Therefore, acidification of surface waters by acidic deposition is not as 

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

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

4 2008, section 3.2.4.2 and US EPA, 2009, section 4.2.2).

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

6 algae that are sensitive to acidification and cannot survive, compete, or reproduce in acidic 

7 waters (US EPA, 2008, section 3.2.3.3). Decreases in ANC and pH have been shown to 

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

23 analyses on aquatic ecosystems in the Adirondack Mountains and Shenandoah National Park 
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1 were conducted to better characterize ecological risk associated with acidification (US EPA, 

2 2009, section 4).

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

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

5 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

-

6 concentrations and ANC levels were 

7 analyzed to affirm the understanding that reductions in deposition could influence the risk of 

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

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

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

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

2 (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.

3

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

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

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

7 pH <6.0 include important recreational species, such as Atlantic salmon, tiger trout (Salmo trutta

8 X Salvelinus fontinalis), redbreast sunfish (Lepomis auritus), bluegill (Lepomis macrochirus), 
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1 tiger musky (Esox masquinongy X lucius), walleye (Sander vitreus), alewife (Alosa 

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

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

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

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

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

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

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

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

declines in NO3

-

and SO4

2-

10 concentrations in surface waters. The 2006 concentrations are still 

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

concentrations of NO3

-

and SO4

2-

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

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

14 (± 4.5 μeq/L). 55% of all monitored streams in the Shenandoah Case Study Area have a current 

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

16 2009, section 4.2.4.3).

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

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

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

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

21 Moderate Concern decreases by 5%.

22 Biological effects of increased acidification documented in the Shenandoah Case Study

23 Area include a decrease in the condition factor in blacknose dace and a decrease in fish 
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1 biodiversity associated with decreasing stream ANC. On average, the fish species richness is 

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

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

4 ii. Terrestrial acidification

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

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

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

(such as Ca2+, Mg2+ 8 and potassium), as well as soil aluminum (Al) content. Tree species show a 

9 range of sensitivities to Ca/Al and Bc/Al soil molar ratios, therefore these are good chemical 

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

11 large variety of tree species ranged from 0.2 to 0.8. This range is similar to critical ratios of 

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

13 from 0.2 to 2.5 (US EPA, 2009).

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

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

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

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

18 >250 eq/ha-yr (McNulty et al., 2007). Forests of the Adirondack Mountains of New York, Green 

19 Mountains of Vermont, White Mountains of New Hampshire, the Allegheny Plateau of 

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

21 most sensitive to terrestrial acidification effects from acidifying deposition (US EPA, 2008, 

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

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

2 time.

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

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

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

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

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

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

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

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

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

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

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

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

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

16 c. Key uncertainties associated with acidification

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

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

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

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

21 relationship between the selected ecological indicator and atmospheric deposition as described 

22 by the ecological effect function (US EPA, 2009, Figure 1-1). A further discussion of 

23 uncertainty in aquatic and terrestrial ecosystems is presented below. The key uncertainties were 
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1 characterized as follows to evaluate the strength of the scientific basis for setting a national 

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

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

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

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

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

7 into this criterion.

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

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

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

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

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

13 atmospheric deposition and the ecological indicator of an effect.

14 i. Aquatic acidification

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

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

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

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

19 1990 to 2006 used for this analysis as well as EPA EMAP and REMAP surveys, provide 

20 considerable data on surface water trends. 

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

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

23 site calibrations. Pre-acidification refers to retrospective modeling to estimate water quality
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1 conditions before man-made contributions of acidifying inputs. The models are evaluated under

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

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

4 concentrations or 10% and 8 μeq/L or 5% for streams (US EPA, 2009, section 7.1.2). The use of 

5 the critical load model to estimate aquatic critical loads is limited by the uncertainties associated 

6 with runoff and surface water measurements and in estimating the catchment supply of base 

7 cations from the weathering of bedrock and soils (McNulty et al., 2007). 

8 ii. Terrestrial acidification 

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

10 robust and considered high quality. The USFS-Kane Experimental Forest and significant 

11 amounts of research work in the Allegheny Plateau have produced extensive, peer-reviewed data 

12 sets. A meta-analysis of laboratory studies showed that tree growth (often measured by changes 

13 in fine root mass) was decreased by 20% relative to controls for BC/Al ratios (US EPA, 2008, 

14 section 7.2.1 and Figure 7.2-1). Sugar maple and red spruce were the focus of the REA since 

they are demonstrated to be negatively affected by soil available Ca2+ 15 depletion and high 

16 concentrations of available Al, and occur in areas that receive high acidifying deposition, There 

17 is high confidence about the use of the REA terrestrial acidification data and their value for 

18 extrapolating to a larger regional population of forests. 

19 There is high confidence associated with the models, input parameters, and assessment of 

20 uncertainty used in the case study for terrestrial acidification. The Simple Mass Balance (SMB) 

21 model, a commonly used and widely applied approach for estimating critical loads, was used in 

22 the REA analysis (US EPA, 2008, section 7.2.2). There is fairly high confidence associated with 
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1 the ecological effect function developed for terrestrial acidification (US EPA, 2009, section

2 7.2.3).

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

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

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

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

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

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

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

10 other sources of N greatly exceed N deposition. Both oxides of nitrogen and reduced forms of 

11 nitrogen (NHx) contribute to nitrogen deposition. For the most part, nitrogen effects on 

12 ecosystems do not depend on whether the nitrogen is in oxidized or reduced form. Thus, this 

13 summary focuses on the effects of nitrogen deposition in total. 

14 a. Nature of nutrient enrichment-related ecosystem responses

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

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

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

18 that leads to effects on biological organisms. 

19 i. Aquatic ecosystems

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

21 N deposition and the alteration of biogeochemical cycling in freshwater aquatic ecosystems (US 

22 EPA, 2008, section 3.3.2.3). N deposition is the main source of N enrichment to headwater 

23 streams, lower order streams and high elevation lakes. The most common chemical indicators 
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that were studied included NO3

−

1 and dissolved inorganic nitrogen (DIN) concentration in surface 

waters as well as Chl a:total P ratio. Elevated surface water NO3

−

2 concentrations occur in both 

3 the eastern and western U.S. Studies report a significant correlation between N deposition and 

4 lake biogeochemistry by identifying a correlation between wet deposition and [DIN] and Chl a: 

5 Total P. Recent evidence provides examples of lakes and streams that are limited by N and show 

6 signs of eutrophication in response to N addition.

7 The evidence is sufficient to infer a causal relationship between N deposition and the 

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

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

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

11 In the ISA, the evidence is sufficient to infer a causal relationship between N deposition 

12 and the biogeochemical cycling of N and carbon (C) in estuaries (US EPA, 2008, section 4.3.4.1 

13 and 3.3.2.3). In general, estuaries tend to be nitrogen-limited, and many currently receive high

14 levels of nitrogen input from human activities (US EPA, 2009, section 5.1.1). It is unknown if 

15 atmospheric deposition alone is sufficient to cause eutrophication; however, the contribution of 

16 atmospheric nitrogen deposition to total nitrogen load is calculated for some estuaries and can be 

17 >40% (US EPA, 2009, section 5.1.1).

18 The evidence is sufficient to infer a causal relationship between N deposition and the 

19 alteration of species richness, species composition and biodiversity in estuarine ecosystems (US

20 EPA, 2008, section 4.3.4.2 and 3.3.5.4). Atmospheric and non-atmospheric sources of N 

21 contribute to increased phytoplankton and algal productivity, leading to eutrophication. Shifts in 

22 community composition, reduced hypolimnetic DO, decreases in biodiversity, and mortality of 

23 submerged aquatic vegetation are associated with increased N deposition in estuarine systems. 
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1 ii. Terrestrial Ecosystems

2 The evidence is sufficient to infer a causal relationship between N deposition and the 

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

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

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

6 of NO3- in stream water were identified as two of the primary indictors of N enrichment. Several 

7 N-addition studies indicate that NO3- leaching is induced by chronic additions of N. Studies 

8 identified in the ISA found that surface water NO3- concentrations exceeded 1 μeq/L in 

9 watersheds receiving about 9 to 13 kg N/ha/yr of atmospheric N deposition. N deposition 

10 disrupts the nutrient balance of ecosystems with numerous biogeochemical effects. The chemical 

indicators that are typically measured include NO3

−

11 leaching, soil C:N ratio, rates of N 

12 mineralization, nitrification, denitrification, foliar N concentration, and soil water NO3 − and 

NH4

+

13 concentrations. Note that N saturation (N leaching from ecosystems) does not need to 

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

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

16 Due to the complexity of interactions between the N and C cycling, the effects of N on C budgets 

17 (quantified input and output of C to the ecosystem) are variable. Regional trends in net 

18 ecosystem productivity (NEP) of forests (not managed for silviculture) have been estimated 

19 through models based on gradient studies and meta-analysis. Atmospheric N deposition has been 

20 shown to cause increased litter accumulation and carbon storage in above-ground woody 

21 biomass. In the West, this has lead to increased susceptibility to more severe fires. Less is 

22 known regarding the effects of N deposition on C budgets of non-forest ecosystems.
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1 The evidence is sufficient to infer a causal relationship between N deposition on the 

2 alteration of species richness, species composition and biodiversity in terrestrial ecosystems (US 

3 EPA, 2008, section 4.3.1.2). Some organisms and ecosystems are more sensitive to N deposition 

4 and effects of N deposition are not observed in all habitats. The most sensitive terrestrial taxa to 

5 N deposition are lichens. Empirical evidence indicates that lichens in the U.S. are affected by

6 deposition levels as low as 3 kg N/ha/yr. Alpine ecosystems are also sensitive to N deposition, 

7 changes in an individual species (Carex rupestris) were estimated to occur at deposition levels 

8 near 4 kg N /ha/yr and modeling indicates that deposition levels near 10 kg N/ha/yr alter plant 

9 community assemblages. In several grassland ecosystems, reduced species diversity and an 

10 increase in non-native, invasive species are associated with N deposition. 

11 iii. Ecosystem sensitivity to nutrient enrichment

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

13 sensitivity to N deposition (US EPA, 2008, Table 4-4). Increased deposition to N-limited 

14 ecosystems can lead to production increases that may be either beneficial or adverse depending 

15 on the system and management goals. 

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

17 nutrient availability. Change in the availability of one important nutrient, such as N, may result 

18 in an imbalance in ecological stoichiometry, with effects on ecosystem processes, structure and 

19 function. In general, N deposition to terrestrial ecosystems causes accelerated growth rates in 

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

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

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

23 ecosystems to N deposition. In aquatic ecosystems, N that is both leached from the soil and 
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1 directly deposited to the water surface can pollute the surface water. This causes alteration of the 

2 diatom community at levels as low as 1.5 kg N/ha/yr in sensitive freshwater ecosystems. 

3 The degree of ecosystem effects lies at the intersection of N loading and N-sensitivity. 

4 N-sensitivity is predominately driven by the degree to which growth is limited by nitrogen 

5 availability. Grasslands in the western United States are typically N-limited ecosystems 

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

7 the ISA (US EPA, 2008, section 3.3.3), indicated that N fertilization increased aboveground 

8 growth in all non-forest ecosystems except for deserts. In other words, almost all terrestrial 

9 ecosystems are N-limited and will be altered by the addition of anthropogenic nitrogen. 

10 Likewise, a freshwater lake or stream must be N-limited to be sensitive to N-mediated 

11 eutrophication. There are many examples of fresh waters that are N-limited or N and 

12 phosphorous (P) co-limited (US EPA, 2008, section 3.3.3.2). A large dataset meta-analysis 

13 discussed in the ISA (US EPA, 2008, section 3.3.3.2), found that N-limitation occurred as 

14 frequently as P-limitation in freshwater ecosystems. Additional factors that govern the 

15 sensitivity of ecosystems to nutrient enrichment from N deposition include rates and form of N 

16 deposition, elevation, climate, species composition, plant growth rate, length of growing season, 

17 and soil N retention capacity (US EPA, 2008, section 4.3). Less is known about the extent and 

18 distribution of the terrestrial ecosystems in the U.S. that are most sensitive to the effects of 

19 nutrient enrichment from atmospheric N deposition compared to acidification.

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

21 limited by the availability of N, they are susceptible to the eutrophication effect of N deposition 

22 (US EPA, 2008, section 4.3.4.1). A recent national assessment of eutrophic conditions in 

23 estuaries found the most eutrophic estuaries were generally those that had large watershed-to-
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1 estuarine surface area, high human population density, high rainfall and runoff, low dilution, and 

2 low flushing rates. In the REA, the National Oceanic and Atmospheric Administration’s 

3 (NOAA) National Estuarine Eutrophication Assessment (NEEA) assessment tool, Assessment of 

4 Estuarine Tropic Status (ASSETS) categorical Eutrophication Index (EI) was used to evaluate 

5 eutrophication due to atmospheric loading of nitrogen. ASSETS EI is an estimation of the 

6 likelihood that an estuary is experiencing eutrophication or will experience eutrophication based 

7 on five ecological indicators: chlorophyll a, macroalgae, dissolved oxygen, nuisance/toxic algal 

8 blooms and submerged aquatic vegetation (SAV). 

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

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

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

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

13 results were then considered with SPAtially Referenced Regression (SPARROW) modeling to 

14 develop a response curve to examine the role of atmospheric nitrogen deposition in achieving a 

15 desired decrease in load. To change the Neuse River Estuary’s EI score from Bad to Poor not 

16 only must 100% of the total atmospheric nitrogen deposition be eliminated, but considerably 

17 more nitrogen from other sources as well must be controlled (US EPA, 2009, section 5.2.7.2). In 

18 the Potomac River estuary, a 78% decrease of total nitrogen could move the EI score from Bad

19 to Poor (US EPA, 2009, section 5.2.7.1). The results of this analysis indicated decreases in 

20 atmospheric deposition alone could not eliminate coastal eutrophication problems due to 

21 multiple non-atmospheric nitrogen inputs (US EPA, 2009, section 7.3.3). However, the 

22 somewhat arbitrary discreteness of the EI scale can mask the benefits of decreases in nitrogen 

23 between categories.
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1 In general, estuaries tend to be N-limited, and many currently receive high levels of N 

2 input from human activities to cause eutrophication. As reported in the ISA (US EPA, 2008, 

3 section 3.2.2.2), atmospheric N loads to estuaries in the U.S. are estimated to range from 2-8% 

4 for Guadalupe Bay, TX on the lowest end to as high as 72% for St Catherines-Sapelo estuary, 

5 GA. The Chesapeake Bay is an example of a large, well-studied and severely eutrophic estuary

6 that is calculated to receive as much as 30% of its total N load from the atmosphere.

7 b. Magnitude of ecosystem responses

8 i. Aquatic ecosystems

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

10 conditions and how ecosystems have been altered since the onset of anthropogenic N deposition. 

11 As noted previously, studies found that N-limitation occurs as frequently as P-limitation in 

12 freshwater ecosystems (US EPA, 2008, section 3.3.3.2). Recently, a comprehensive study of 

13 available data from the northern hemisphere surveys of lakes along gradients of N deposition 

14 show increased inorganic N concentration and productivity to be correlated with atmospheric N 

15 deposition. The results are unequivocal evidence of N limitation in lakes with low ambient inputs 

16 of N, and increased N concentrations in lakes receiving N solely from atmospheric N deposition. 

17 It has been suggested that most lakes in the northern hemisphere may have originally been N18 limited, and that atmospheric N deposition has changed the balance of N and P in lakes.

19 Available data suggest that the increases in total N deposition do not have to be large to 

20 elicit an ecological effect. For example, a hindcasting exercise determined that the change in 

21 Rocky Mountain National Park lake algae that occurred between 1850 and 1964 was associated 

22 with an increase in wet N deposition that was only about 1.5 kg N/ha. Similar changes inferred 

23 from lake sediment cores of the Beartooth Mountains of Wyoming also occurred at about 1.5 kg 
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1 N/ha deposition. Pre-industrial inorganic N deposition is estimated to have been only 0.1 to 0.7 

2 kg N/ha based on measurements from remote parts of the world. In the western U.S., pre3 industrial, or background, inorganic N deposition was estimated by to range from 0.4 to 0.7 kg 

4 N/ha/yr.

5 Eutrophication effects from N deposition are most likely to be manifested in undisturbed, 

6 low nutrient surface waters such as those found in the higher elevation areas of the western U.S. 

7 The most severe eutrophication from N deposition effects is expected downwind of major urban 

and agricultural centers. High concentrations of lake or streamwater NO3

−

8 , indicative of 

9 ecosystem saturation, have been found at a variety of locations throughout the U.S., including the 

10 San Bernardino and San Gabriel Mountains within the Los Angeles Air Basin, the Front Range 

11 of Colorado, the Allegheny mountains of West Virginia, the Catskill Mountains of New York, 

12 the Adirondack Mountains of New York, and the Great Smoky Mountains in Tennessee (US 

13 EPA, 2008, section 3.3.8).

14 In contrast to terrestrial and freshwater systems, atmospheric N load to estuaries 

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

16 sources of N often greatly exceed N deposition. In estuaries, N-loading from multiple 

17 anthropogenic and non-anthropogenic pathways leads to water quality deterioration, resulting in 

18 numerous effects including hypoxic zones, species mortality, changes in community composition 

19 and harmful algal blooms that are indicative of eutrophication. The following summary is a 

20 concise overview of the known or anticipated effects of nitrogen enrichment on estuaries within 

21 the United States.

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

23 eutrophication in shallow estuaries has increased over the past several decades and that the 
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1 environmental degradation of coastal ecosystems due to nitrogen, phosphorus, and other inputs is 

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

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

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

5 rivers’ discharges to the Gulf of Mexico. 

6 A recent national assessment of eutrophic conditions in estuaries found that 65% of the 

7 assessed systems had moderate to high overall eutrophic conditions. Most eutrophic estuaries 

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

9 were in the North Atlantic. Other regions had mixtures of low, moderate, and high degrees of 

10 eutrophication (US EPA, 2008, section 4.3.4.3).

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

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

13 Submerged aquatic vegetation is important to the quality of estuarine ecosystem habitats because 

14 it provides habitat for a variety of aquatic organisms, absorbs excess nutrients, and traps 

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

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

17 negative impacts from nitrogen addition from atmospheric deposition.

18 ii. Terrestrial ecosystems

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

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

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

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

23 likely to occur where areas of relatively high atmospheric N deposition intersect with N-limited 
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1 plant communities. The alpine ecosystems of the Colorado Front Range, chaparral watersheds of 

2 the Sierra Nevada, lichen and vascular plant communities in the San Bernardino Mountains and 

3 the Pacific Northwest, and the southern California coastal sage scrub (CSS) community are 

4 among the most sensitive terrestrial ecosystems. There is growing evidence (US EPA, 2008, 

5 section 4.3.1.2) that existing grassland ecosystems in the western United States are being altered 

6 by elevated levels of N inputs, including inputs from atmospheric deposition.

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

assessed in terms of the degree of NO3

−

8 leaching from watershed soils into ground water or 

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

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

11 60% were in Stage 1 or Stage 2 of N saturation on a scale of 0 (background or pretreatment) to 3 

12 (visible decline). Of the northeastern sites for which adequate data were available for assessment, 

13 those in Stage 1 or 2 were most prevalent in the Adirondack and Catskill Mountains. Effects on 

14 individual plant species have not been well studied in the U.S. More is known about the 

15 sensitivity of particular plant communities. Based largely on results obtained in more extensive 

16 studies conducted in Europe, it is expected that the more sensitive terrestrial ecosystems include 

17 hardwood forests, alpine meadows, arid and semi-arid lands, and grassland ecosystems (US 

18 EPA, 2008, section 3.3.5).

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

20 2008, Table 4.4) to identify meaningful ecological benchmarks associated with different levels of 

21 atmospheric nitrogen deposition. These are illustrated in Figure 3-4 of the Policy Assessment. 

22 The sensitive areas and ecological indicators identified by the ISA were analyzed further in the 

23 REA to create a national map that illustrates effects observed from ambient and experimental 
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1 atmospheric nitrogen deposition loads in relation to Community Multi-scale Air Quality 

2 (CMAQ) 2002 modeling results and NADP monitoring data. This map, reproduced in Figure 3-

3 5 of the Policy Assessment, depicts the sites where empirical effects of terrestrial nutrient 

4 enrichment have been observed and site proximity to elevated atmospheric N deposition. 

5 Based on information in the ISA and initial analysis in the REA, further case study

6 analyses on terrestrial nutrient enrichment of ecosystems were developed for the CCS 

7 community and Mixed Conifer Forest (MCF) (US EPA, 2009). Geographic information systems 

8 (GIS) analysis supported a qualitative review of past field research to identify ecological 

9 benchmarks associated with CSS and mycorrhizal communities, as well as MCF nutrient10 sensitive acidophyte lichen communities, fine-root biomass in Ponderosa pine, and leached 

11 nitrate in receiving waters. 

12 The ecological benchmarks that were identified for the CSS and the MCF communities

13 are included in the suite of benchmarks identified in the ISA (US EPA, 2008, section 3.3). There 

14 are sufficient data to confidently relate the ecological effect to a loading of atmospheric nitrogen. 

15 For the CSS community, the following ecological benchmarks were identified:

16 (1) 3.3 kg N/ha/yr – the amount of nitrogen uptake by a vigorous stand of CSS; above this level, 

17 nitrogen may no longer be limiting

18 (2) 10 kg N/ha/yr – mycorrhizal community changes

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

20 (1) 3.1 kg N/ha/yr – shift from sensitive to tolerant lichen species

21 (2) 5.2 kg N/ha/yr – dominance of the tolerant lichen species

22 (3) 10.2 kg N/ha/yr – loss of sensitive lichen species

23 (4) 17 kg N/ha/yr – leaching of nitrate into streams.
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1 These benchmarks, ranging from 3.1 to 17 kg N/ha/yr, were compared to 2002 

2 CMAQ/NADP data to discern any associations between atmospheric deposition and changing 

3 communities. Evidence supports the finding that nitrogen alters CSS and MCF communities. 

4 Key findings include the following: 2002 CMAQ/NADP nitrogen deposition data show that the 

5 3.3 kg N/ha/yr benchmark has been exceeded in more than 93% of CSS areas (654,048 ha). 

6 These deposition levels are a driving force in the degradation of CSS communities. Although 

7 CSS decline has been observed in the absence of fire, the contributions of deposition and fire to 

8 the CSS decline require further research. CSS is fragmented into many small parcels, and the 

9 2002 CMAQ/NADP 12-km grid data are not fine enough to fully validate the relationship 

10 between CSS distribution, nitrogen deposition, and fire. 2002 CMAQ/NADP nitrogen deposition 

11 data exceeds the 3.1 kg N/ha/yr benchmark in more than 38% (1,099,133 ha) of MCF areas, and 

12 nitrate leaching has been observed in surface waters. Ozone effects confound nitrogen effects on 

13 MCF acidophyte lichen, and the interrelationship between fire and nitrogen cycling requires 

14 additional research.

15 c. Key uncertainties associated with nutrient enrichment

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

17 ecological effects and ecological indicators. The criteria used in the REA to evaluate the degree 

18 of confidence in the data, modeling and ecological effect function are detailed in Chapter 7 of the 

19 REA. Below is a discussion of uncertainty relating aquatic and terrestrial ecosystems to nutrient 

20 enrichment effects.

21 i. Aquatic ecosystems 

22 The approach for assessing atmospheric contributions to total nitrogen loading in the 

23 REA was to consider the main-stem river to an estuary (including the estuary) rather than an 
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1 entire estuary system or bay. The biological indicators used in the NOAA ASSETS EI required 

2 the evaluation of many national databases including the US Geological Survey National Water 

3 Quality Assessment (NAWQA) files, EPA’s STORage and RETrieval (STORET) database, 

4 NOAA’s Estuarine Drainage Areas data, and EPA’s water quality standards nutrient criteria for 

5 rivers and lakes (US EPA, 2009, Appendix 6 and Table 1.2.-1). Both the SPARROW modeling 

6 for nitrogen loads and assessment of estuary conditions under NOAA ASSETS EI, have been 

7 applied on a national scale. The REA concludes that the available data are medium quality with 

8 intermediate confidence about the use of these data and their values for extrapolating to a larger 

9 regional area (US EPA, 2009, section 7.3.1). Intermediate confidence is associated with the 

10 modeling approach using ASSETS EI and SPARROW. The REA states there is low confidence 

11 with the ecological effect function due to the results of the analysis which indicated that 

12 reductions in atmospheric deposition alone could not solve coastal eutrophication problems due 

13 to multiple non-atmospheric nitrogen inputs (US EPA, 2009, section 7.3.3).

14 ii. Terrestrial ecosystems

15 Ecological thresholds are identified for CSS and MCF areas and these data are considered 

16 to be of high quality, however, the ability to extrapolate these data to larger regional areas is 

17 limited (US EPA, 2009, section 7.4.1). No quantitative modeling was conducted or ecological 

18 effect function developed for terrestrial nutrient enrichment reflecting the uncertainties 

19 associated with these depositional effects. 

20 4. Other Ecological Effects 

21 It is stated in the ISA (US EPA, 2008, section 3.4.1 and 4.5) that mercury is a highly 

22 neurotoxic contaminant that enters the food web as a methylated compound, methylmercury. 

23 Mercury is principally methylated by sulfur-reducing bacteria and can be taken up by
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1 microorganisms, zooplankton and macroinvertebrates. The contaminant is concentrated in higher 

2 trophic levels, including fish eaten by humans. Experimental evidence has established that only 

3 inconsequential amounts of methylmercury can be produced in the absence of sulfate. Once 

4 methylmercury is present, other variables influence how much accumulates in fish, but elevated 

5 mercury levels in fish can only occur where substantial amounts of methylmercury are present. 

6 Current evidence indicates that in watersheds where mercury is present, increased oxides of 

7 sulfur deposition very likely results in additional production of methylmercury which leads to 

8 greater accumulation of MeHg concentrations in fish. With respect to sulfur deposition and 

9 mercury methylation, the final ISA determined: The evidence is sufficient to infer a causal 

10 relationship between sulfur deposition and increased mercury methylation in wetlands and 

11 aquatic environments. 

12 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-

13 results in 

14 increased production of MeHg. There is increasing evidence on the relationship between sulfur 

15 deposition and increased methylation of mercury in aquatic environments; this effect occurs only 

16 where other factors are present at levels within a range to allow methylation. The production of 

17 methylmercury requires the presence of sulfate and mercury, but the amount of methylmercury

18 produced varies with oxygen content, temperature, pH, and supply of labile organic carbon (US 

19 EPA, 2008, section 3.4). In watersheds where changes in sulfate deposition did not produce an 

20 effect, one or several of those interacting factors were not in the range required for meaningful 

21 methylation to occur (US EPA, 2008, section 3.4). Watersheds with conditions known to be 

22 conducive to mercury methylation can be found in the northeastern United States and 

23 southeastern Canada. 
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1 While the relationship between sulfur and methylmercury production was concluded to 

2 be causal in the ISA, the REA concluded that there was insufficient evidence to quantify the

3 relationship between sulfur and methylmercury. Therefore only a qualitative assessment was 

4 included in Chapter 6 of the REA. The Policy Assessment was then unable to make a 

5 determination as to the adequacy of the existing SO2 standards in protecting against welfare 

6 effects associated with increased mercury methylation.

7 B. Risk and Exposure Assessment 

8 The risk and exposure assessment conducted for the current review was developed to 

9 describe potential risk from current and future deposition of oxides of nitrogen and sulfur to 

10 sensitive ecosystems. The case study analyses in the REA show that there is confidence that 

11 known or anticipated adverse ecological effects are occurring under current ambient loadings of 

12 nitrogen and sulfur in sensitive ecosystems across the United States. An overview of the 

13 material covered in the REA, a summary of the key findings from the air quality analyses, 

14 acidification and nutrient enrichment case studies, and general conclusions from evaluating 

15 additional welfare effects, are presented below.

16 1. Overview of the Risk and Exposure Assessment

17 The REA evaluates the relationships between atmospheric concentrations, deposition, 

18 biologically relevant exposures, targeted ecosystem effects, and ecosystem services. To evaluate 

19 the nature and magnitude of adverse effects associated with deposition, the REA also examines 

20 various ways to quantify the relationships between air quality indicators, deposition of 

21 biologically available forms of nitrogen and sulfur, ecologically relevant indicators relating to 

22 deposition, exposure and effects on sensitive receptors, and related effects resulting in changes in 

23 ecosystem structure and services. The intent is to determine the exposure metrics that incorporate 
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1 the temporal considerations (i.e., biologically relevant timescales), pathways, and ecologically 

2 relevant indicators necessary to determine the effects on these ecosystems. To the extent feasible, 

3 the REA evaluates the overall load to the system for nitrogen and sulfur, as well as the variability

4 in ecosystem responses to these pollutants. It also evaluates the contributions of atmospherically 

5 deposited nitrogen and sulfur individually relative to the combined atmospheric loadings of both 

6 elements together. Since oxidized nitrogen is the listed criteria pollutant (currently measured by 

7 the ambient air quality indicator NO2) for the atmospheric contribution to total nitrogen, the REA 

8 examines the contribution of nitrogen oxides to total reactive nitrogen in the atmosphere, relative 

9 to the contributions of reduced forms of nitrogen (e.g., ammonia, ammonium), to ultimately

10 assess how a meaningful secondary National Ambient Air Quality Standards (NAAQS) might be 

11 structured.

12 The REA focuses on ecosystem welfare effects that result from the deposition of total 

13 reactive nitrogen and sulfur. Because ecosystems are diverse in biota, climate, geochemistry, and 

14 hydrology, response to pollutant exposures can vary greatly between ecosystems. In addition, 

15 these diverse ecosystems are not distributed evenly across the United States. To target nitrogen 

16 and sulfur acidification and nitrogen and sulfur enrichment, the REA addresses four main 

17 targeted ecosystem effects on terrestrial and aquatic systems identified by the ISA (US EPA, 

18 2008): aquatic acidification due to nitrogen and sulfur; terrestrial acidification due to nitrogen 

19 and sulfur; aquatic nutrient enrichment, including eutrophication; and terrestrial nutrient 

20 enrichment.

21 In addition to these four targeted ecosystem effects, the REA also qualitatively addresses 

22 the influence of sulfur oxides deposition on methylmercury production; nitrous oxide (N2O) 
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1 effects on climate; nitrogen effects on primary productivity and biogenic greenhouse gas fluxes; 

2 and phytotoxic effects on plants. 

3 Because the targeted ecosystem effects outlined above are not evenly distributed across 

4 the United States, the REA identified case studies for each targeted effects based on ecosystems 

5 identified as sensitive to nitrogen and/or sulfur deposition effects. Eight case study areas and 

6 two supplemental study areas (Rocky Mountain National Park and Little Rock Lake, WI) are 

7 summarized in the REA based on ecosystem characteristics, indicators, and ecosystem service 

8 information. Case studies selected for aquatic acidification effects were the Adirondack 

9 Mountains and Shenandoah National Park. Kane Experimental Forest in Pennsylvania and 

10 Hubbard Brook Experimental Forest in New Hampshire were selected as case studies for 

11 terrestrial acidification. Aquatic nutrient enrichment case study locations were selected in the 

12 Potomac River Basin upstream of Chesapeake Bay and the Neuse River Basin upstream of the 

13 Pamlico Sound in North Carolina. The Coastal Sage Scrub Communities in southern California 

14 and the Mixed Confer Forest Communities in the San Bernardino and Sierra Nevada Mountains 

15 of California were selected as case studies for terrestrial nutrient enrichment. Two supplemental 

16 areas were also chosen, one in Rocky Mountain National Park for terrestrial nutrient enrichment 

17 and one in Little Rock Lake, Wisconsin for aquatic nutrient enrichment.

18 2. Key findings

19 In summary, based on case study analyses, the REA concludes that known or anticipated 

20 adverse ecological effects are occurring under current conditions and further concludes that these 

21 adverse effects continue into the future. Key findings from the air quality analyses, acidification 

22 and nutrient enrichment case studies, as well as general conclusions from evaluating additional 

23 welfare effects, are summarized below.
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1 a. Air quality analyses

2 The air quality analyses in the REA encompass the current emissions sources of nitrogen 

3 and sulfur, as well as atmospheric concentrations, estimates of deposition of total nitrogen, 

4 policy-relevant background, and non-atmospheric loadings of nitrogen and sulfur to ecosystems, 

5 both nationwide and in the case study areas. Spatial fields of deposition were created using wet 

6 deposition measurements from the National Atmospheric Deposition Program (NADP) National 

7 Trends Network and dry deposition predictions from the 2002 Community Multi-Scale Air 

8 Quality (CMAQ) model simulation. Some key conclusions from this analysis are:

9 (1) Total reactive nitrogen deposition and sulfur deposition are much greater in the East 

10 compared to most areas of the West.

11 (2) These regional differences in deposition correspond to the regional differences in oxides of 

12 nitrogen and SO2 concentrations and emissions, which are also higher in the East. Oxides of 

13 nitrogen emissions are much greater and generally more widespread than NH3 emissions 

14 nationwide; high NH3 emissions tend to be more local (e.g., eastern North Carolina) or sub15 regional (e.g., the upper Midwest and Plains states). The relative amounts of oxidized versus 

16 reduced nitrogen deposition are consistent with the relative amounts of oxides of nitrogen

17 and NH3 emissions. Oxidized nitrogen deposition exceeds reduced nitrogen deposition in 

18 most of the case study areas; the major exception being the Neuse River/Neuse River Estuary

19 Case Study Area.

20 (3) Reduced nitrogen deposition exceeds oxidized nitrogen deposition in the vicinity of local 

21 sources of NH3.
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1 (4) There can be relatively large spatial variations in both total reactive nitrogen deposition and 

2 sulfur deposition within a case study area; this occurs particularly in those areas that contain 

3 or are near a high emissions source of oxides of nitrogen, NH3, and/or SO2.

4 (5) The seasonal patterns in deposition differ between the case study areas. For the case study 

5 areas in the East, the season with the greatest amounts of total reactive nitrogen deposition 

6 correspond to the season with the greatest amounts of sulfur deposition. Deposition peaks in 

7 spring in the Adirondack, Hubbard Brook Experimental Forest, and Kane Experimental 

8 Forest case study areas, and it peaks in summer in the Potomac River/Potomac Estuary, 

9 Shenandoah, and Neuse River/Neuse River Estuary case study areas. For the case study areas 

10 in the West, there is less consistency in the seasons with greatest total reactive nitrogen and 

11 sulfur deposition in a given area. In general, both nitrogen and/or sulfur deposition peaks in 

12 spring or summer. The exception to this is the Sierra Nevada Range portion of the Mixed 

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

14 b. Deposition-related aquatic acidification

15 The role of aquatic acidification in two eastern United States areas—northeastern New 

16 York’s Adirondack area and the Shenandoah area in Virginia—was analyzed in the REA to 

assess surface water trends in SO4

2-

and NO3

-

17 concentrations and acid neutralizing capacity (ANC) 

18 levels and to affirm the understanding that reductions in deposition could influence the risk of 

19 acidification. Monitoring data from the EPA-administered Temporally Integrated Monitoring of 

20 Ecosystems (TIME)/Long-Term Monitoring (LTM) programs and the Environmental Monitoring 

21 and Assessment Program (EMAP) were assessed for the years 1990 to 2006, and past, present, 

22 and future water quality levels were estimated using both steady-state and dynamic 

23 biogeochemical models. 
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1 Although wet deposition rates for SO2 and oxides of nitrogen in the Adirondack Case 

2 Study Area have reduced since the mid-1990s, current concentrations are still well above pre3 acidification (1860) conditions. Model of Acidification of Groundwater in Catchments (MAGIC) 

modeling predicts NO3

-

and SO4

2-

4 are 17- and 5-fold higher today, respectively. The estimated 

5 average ANC for 44 lakes in the Adirondack Case Study Area is 62.1 microequivalents per liter 

6 (μeq/L) (± 15.7 μeq/L); 78 % of all monitored lakes in the Adirondack Case Study Area have a 

7 current risk of Elevated, Severe, or Acute. Of the 78%, 31% experience episodic acidification, 

8 and 18% are chronically acidic today.

9 (1) Based on the steady-state critical load model for the year 2002, 18%, 28%, 44%, and 58% of 

10 169 modeled lakes received combined total sulfur and nitrogen deposition that exceeded 

11 critical loads corresponding to ANC limits of 0, 20, 50, and 100 μeq/L respectively.

12 (2) Based on a deposition scenario that maintains current emission levels to 2020 and 2050, the 

13 simulation forecast indicates no improvement in water quality in the Adirondack Case Study 

14 Area. The percentage of lakes within the Elevated to Acute Concern classes remains the same 

15 in 2020 and 2050.

16 (3) Since the mid-1990s, streams in the Shenandoah Case Study Area have shown slight declines 

in NO3 and SO4

2-

17 concentrations in surface waters. ANC levels increased from about 50 

18 μeq/L in the early 1990s to >75 μeq/L until 2002, when ANC levels declined back to 1991–

19 1992 levels. Current concentrations are still above pre-acidification (1860) conditions. 

MAGIC modeling predicts surface water concentrations of NO3 and SO4

2-

20 are 10- and 32-

21 fold higher today, respectively. The estimated average ANC for 60 streams in the 

22 Shenandoah Case Study Area is 57.9 μeq/L (± 4.5 μeq/L). 55% of all monitored streams in 
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1 the Shenandoah Case Study Area have a current risk of Elevated, Severe, or Acute. Of the 

2 55%, 18% experience episodic acidification, and 18% are chronically acidic today.

3 (4) Based on the steady-state critical load model for the year 2002, 52%, 72%, 85%, and 93% of 

4 60 modeled streams received combined total sulfur and nitrogen deposition that exceeded 

5 critical loads corresponding to ANC limits of 0, 20, 50, and 100 μeq/L respectively.

6 (5) Based on a deposition scenario that maintains current emission levels to 2020 and 2050, the 

7 simulation forecast indicates that a large number of streams would still have Elevated to 

8 Acute problems with acidity. In fact, from 2006 to 2050, the percentage of streams with 

9 Acute Concern increases by 5%, while the percentage of streams in Moderate Concern 

10 decreases by 5%.

11 c. Deposition-related terrestrial acidification

12 The role of terrestrial acidification was examined in the REA using a critical load 

13 analysis for sugar maple and red spruce forests in the eastern United States by using the base 

14 cation to aluminum (Bc/Al) ratio in acidified forest soils as an indicator to assess the impact of 

15 nitrogen and sulfur deposition on tree health. These are the two most commonly studied species 

16 in North America for impacts of acidification. At a Bc/Al ratio of 1.2, red spruce growth can be 

17 reduced by 20%. Sugar maple growth can be reduced by 20% at a Bc/Al ratio of 0.6. Key 

18 findings of the case study are summarized below.

19 (1) Case study results suggest that the health of at least a portion of the sugar maple and red 

20 spruce growing in the United States may have been compromised with acidifying total 

21 nitrogen and sulfur deposition in 2002. 2002 CMAQ/NADP total nitrogen and sulfur 

22 deposition levels exceeded three selected critical loads in 3% to 75% of all sugar maple plots 

23 across 24 states. The three critical loads ranged from 6,008 to 107 eq/ha/yr for the Bc/Al 
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1 ratios of 0.6, 1.2, and 10.0 (increasing levels of tree protection). The 2002 CMAQ/NADP 

2 total nitrogen and sulfur deposition levels exceeded three selected critical loads in 3% to 36% 

3 of all red spruce plots across 8 states. The three critical loads ranged from 4,278 to 180 

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

5 (2) The Simple Mass Balance model assumptions made for base cation weathering (Bcw) and 

6 forest soil ANC input parameters are the main sources of uncertainty since these parameters 

7 are rarely measured and require researchers to use default values..

8 (3) The pattern of case study results suggests that nitrogen and sulfur acidifying deposition in the 

9 sugar maple and red spruce forest areas studied were similar in magnitude to the critical 

10 loads for those areas and both ecosystems are likely to be sensitive to any future changes in 

11 the levels of deposition.

12 d. Deposition-related aquatic nutrient enrichment

13 The role of nitrogen deposition in two main stem rivers feeding their respective estuaries 

14 was analyzed in the REA to determine if decreases in deposition could influence the risk of 

15 eutrophication as predicted using the Assessment of Estuarine Trophic Status eutrophication 

16 index (ASSETS EI) scoring system in tandem with SPAtially Referenced Regression on 

17 Watershed Attributes (SPARROW) modeling. This modeling approach provides a transferrable, 

18 intermediate-level analysis of the linkages between atmospheric deposition and receiving waters, 

19 while providing results on which conclusions could be drawn. A summary of findings follows:

20 (1) 2002 CMAQ/NADP results showed that an estimated 40,770,000 kg of total nitrogen was 

21 deposited in the Potomac River watershed. SPARROW modeling predicted that 7,380,000 kg 

22 N/yr of the deposited nitrogen reached the estuary (20% of the total load to the estuary). The 
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1 overall ASSETS EI for the Potomac River and Potomac Estuary was Bad (based on all 

2 sources of N).

3 (2) To improve the Potomac River and Potomac Estuary ASSETS EI score from Bad to Poor, a 

4 decrease of at least 78% in the 2002 total nitrogen atmospheric deposition load to the 

5 watershed would be required.

6 (3) 2002 CMAQ/NADP results showed that an estimated 18,340,000 kg of total nitrogen was 

7 deposited in the Neuse River watershed. SPARROW modeling predicted that 1,150,000 kg 

8 N/yr of the deposited nitrogen reached the estuary (26% of the total load to the estuary). The 

9 overall ASSETS EI for the Neuse River/Neuse River Estuary was Bad.

10 (4) It was found that the Neuse River/Neuse River Estuary ASSETS EI score could not be 

11 improved from Bad to Poor with decreases only in the 2002 atmospheric deposition load to 

12 the watershed. Additional reductions would be required from other nitrogen sources within 

13 the watershed.

14 The small effect of decreasing atmospheric deposition in the Neuse River watershed is 

15 because the other nitrogen sources within the watershed are more influential than atmospheric 

16 deposition in affecting the total nitrogen loadings to the Neuse River Estuary, as estimated with 

17 the SPARROW model. A waterbody’s response to nutrient loading depends on the magnitude 

18 (e.g., agricultural sources have a high influence in the Neuse than in the Potomac), spatial 

19 distribution, and other characteristics of the sources within the watershed.

20 e. Deposition-related terrestrial nutrient enrichment

21 California Coastal Sage Scrub (CSS) and Mixed Conifer Forest (MCF) communities 

22 were the focus of the Terrestrial Nutrient Enrichment Case Studies of the REA. Geographic 

23 information systems (GIS) analysis supported a qualitative review of past field research to 
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1 identify ecological benchmarks associated with CSS and mycorrhizal communities, as well as 

2 MCF’s nutrient-sensitive acidophyte lichen communities, fine-root biomass in Ponderosa pine, 

3 and leached nitrate in receiving waters. These benchmarks, ranging from 3.1 to 17 kg N/ha/yr, 

4 were compared to 2002 CMAQ/NADP data to discern any associations between atmospheric 

5 deposition and changing communities. Evidence supports the finding that nitrogen alters CSS 

6 and MCF. Key findings include the following:

7 (1) 2002 CMAQ/NADP nitrogen deposition data show that the 3.3 kg N/ha/yr benchmark has 

8 been exceeded in more than 93% of CSS areas (654,048 ha). These deposition levels are a 

9 driving force in the degradation of CSS communities. Although CSS decline has been 

10 observed in the absence of fire, the contributions of deposition and fire to the CSS decline 

11 require further research. CSS is fragmented into many small parcels, and the 2002 

12 CMAQ/NADP 12-km grid data are not fine enough to fully validate the relationship between 

13 CSS distribution, nitrogen deposition, and fire.

14 (2) 2002 CMAQ/NADP nitrogen deposition data exceeds the 3.1 kg N/ha/yr benchmark in more 

15 than 38% (1,099,133 ha) of MCF areas, and nitrate leaching has been observed in surface 

16 waters. Ozone effects confound nitrogen effects on MCF acidophyte lichen, and the 

17 interrelationship between fire and nitrogen cycling requires additional research.

18 f. Additional effects

19 Ecological effects have also been documented across the United States where elevated 

20 nitrogen deposition has been observed, including the eastern slope of the Rocky Mountains 

21 where shifts in dominant algal species in alpine lakes have occurred where wet nitrogen 

22 deposition was only about 1.5 kg N/ha/yr. High alpine terrestrial communities have a low 
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1 capacity to sequester nitrogen deposition, and monitored deposition exceeding 3 to 4 kg N/ha/yr 

2 could lead to community-level changes in plant species, lichens, and mycorrhizae.

3 Additional welfare effects are documented, but examined less extensively, in the REA. 

4 These effects include qualitative discussions related to visibility and materials damage, such as 

5 corrosion, erosion, and soiling of paint and buildings which are being addressed in the particulate 

6 matter (PM) NAAQS review currently underway. A discussion of the causal relationship 

between sulfur deposition (as sulfate, SO4

2-

7 ) and increased mercury methylation in wetlands and 

8 aquatic environments is also included in the REA. On this subject the REA concludes that 

9 decreases in sulfate deposition will likely result in decreases in methyl mercury concentration; 

10 however, spatial and biogeochemical variations nationally hinder establishing large scale dose11 response relationships. 

12 Several additional issues concerning oxides of nitrogen were addressed in the REA. 

13 Consideration was also given to nitrous oxide (N2O), a potent GHG. The REA concluded that it 

14 is most appropriate to analyze the role of N2O in the context of all of the GHGs rather than as 

15 part of the REA for this review. The REA considered nitrogen deposition and its correlation 

16 with the rate of photosynthesis and net primary productivity. Nitrogen addition ranging from 

17 15.4 to 300 kg N/ha/yr is documented as increasing wetland N2O production by an average of 

18 207% across all ecosystems. Nitrogen addition ranging from 30 to 240 kg N/ha/yr increased CH4

19 emissions by 115%, averaged across all ecosystems, and methane uptake was reduced by 38%

20 across all ecosystems when nitrogen addition ranged from 10 to 560 kg N/ha/yr, but reductions 

21 were only significant for coniferous and deciduous forests. The heterogeneity of ecosystems 

22 across the United States, however, introduces variations into dose-response relationships.
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1 The phytotoxic effects of oxides of nitrogen and sulfur on vegetation were also briefly 

2 discussed in the REA which concluded that since a unique secondary NAAQS exists for SO2, 

3 and concentrations of nitric oxide (NO), NO2, and PAN are rarely high enough to have 

4 phytotoxic effects on vegetation, further assessment was not warranted at this time. 

5 3. Conclusions on Effects 

6 For aquatic and terrestrial acidification effects, a similar conceptual approach was used 

7 (critical loads) to evaluate the impacts of multiple pollutants on an ecological endpoint, whereas 

8 the approaches used for aquatic and terrestrial nutrient enrichment were fundamentally distinct. 

9 Although the ecological indicators for aquatic and terrestrial acidification (i.e., ANC and Bc/Al) 

10 are very different, both ecological indicators are well-correlated with effects such as reduced 

11 biodiversity and growth. While aquatic acidification is clearly the targeted effect area with the 

12 highest level of confidence, the relationship between atmospheric deposition and an ecological 

13 indicator is also quite strong for terrestrial acidification. The main drawback with the 

14 understanding of terrestrial acidification is that the data are based on laboratory responses rather 

15 than field measurements. Other stressors that are present in the field but that are not present in 

16 the laboratory may confound this relationship.

17 For nutrient enrichment effects, the REA utilized different types of indicators for aquatic 

18 and terrestrial effects to assess both the likelihood of adverse effects to ecosystems and the 

19 relationship between adverse effects and atmospheric sources of oxides of nitrogen. The 

20 ecological indicator chosen for aquatic nutrient enrichment, the ASSETS EI, seems to be 

21 inadequate to relate atmospheric deposition to the targeted ecological effect, likely due to the 

22 many other confounding factors. Further, there is far less confidence associated with the 

23 understanding of aquatic nutrient enrichment because of the large contributions from non-
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1 atmospheric sources of nitrogen and the influence of both oxidized and reduced forms of 

2 nitrogen, particularly in large watersheds and coastal areas. However, a strong relationship exists 

3 between atmospheric deposition of nitrogen and ecological effects in high alpine lakes in the 

4 Rocky Mountains because atmospheric deposition is the only source of nitrogen to these 

5 systems. There is also a strong weight-of-evidence regarding the relationships between 

6 ecological effects attributable to terrestrial nitrogen nutrient enrichment; however, ozone and 

7 climate change may be confounding factors. In addition, the response for other species or species 

8 in other regions of the United States has not been quantified.

9 C. Adversity of Effects to Public Welfare

10 Characterizing a known or anticipated adverse effect to public welfare is an important 

11 component of developing any secondary NAAQS. According to the Clean Air Act, welfare 

12 effects include: “effects on soils, water, crops, vegetation, manmade materials, animals, wildlife, 

13 weather, visibility, and climate, damage to and deterioration of property, and hazards to 

14 transportation, as well as effect on economic values and on personal comfort and well-being, 

15 whether caused by transformation, conversion, or combination with other air pollutants” (CAA, 

16 Section 302(h)). While the text above lists a number of welfare effects, these effects do not 

17 define public welfare in and of themselves. 

18 Although there is no specific definition of adversity to public welfare, the paradigm of 

19 linking adversity to public welfare to disruptions in ecosystem structure and function has been 

20 used broadly by EPA to categorize effects of pollutants from the cellular to the ecosystem level. 

21 An evaluation of adversity to public welfare might consider the likelihood, type, magnitude, and 

22 spatial scale of the effect as well as the potential for recovery and any uncertainties relating to 

23 these considerations. 
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1 Similar concepts were used in past reviews of secondary NAAQS for ozone and PM 

2 (relating to visibility), as well as in initial reviews of effects from lead deposition. Because 

3 oxides of nitrogen and sulfur are deposited from ambient sources into ecosystems where they

4 affect changes to organisms, populations and ecosystems, the concept of adversity to public 

5 welfare as a result of alterations in structure and function of ecosystems is an appropriate 

6 consideration for this review. 

7 Based on information provided in the Policy Assessment, the following section discusses 

8 how ecological effects from deposition of oxides of nitrogen and sulfur relate to adversity to 

9 public welfare. In the Policy Assessment, public welfare was discussed in terms of loss of 

10 ecosystem services (defined below), which in some cases can be monetized. Each of the four 

11 main effect areas (aquatic and terrestrial acidification and aquatic and terrestrial nutrient over12 enrichment) are discussed including current ecological effects and associated ecosystem services.

13 1. Ecosystem Services 

14 The Policy Assessment defines ecosystem services as the benefits individuals and 

15 organizations obtain from ecosystems. Ecosystem services can be classified as provisioning 

16 (food and water), regulating (control of climate and disease), cultural (recreational, existence, 

17 spiritual, educational), and supporting (nutrient cycling). Conceptually, changes in ecosystem 

18 services may be used to aid in characterizing a known or anticipated adverse effect to public 

19 welfare. In the REA and Policy Assessment ecosystem services are discussed as a method of 

20 assessing the magnitude and significance to the public of resources affected by ambient 

21 concentrations of oxides of nitrogen and sulfur and deposition in sensitive ecosystems. 

22 EPA has in previous NAAQS reviews defined ecological goods and services for the 

23 purposes of a Regulatory Impact Analysis as the “outputs of ecological functions or processes 
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1 that directly or indirectly contribute to social welfare or have the potential to do so in the future. 

2 Some outputs may be bought and sold, but most are not marketed.” It is especially important to

3 acknowledge that it is difficult to measure and/or monetize the goods and services supplied by

4 ecosystems. It can be informative in characterizing adversity to public welfare to attempt to 

5 place an economic valuation on the set of goods and services that have been identified with 

6 respect to a change in policy however it must be noted that this valuation will be incomplete and 

7 illustrative only. 

8 Knowledge about the relationships linking ambient concentrations and ecosystem 

9 services is considered in the Policy Assessment as one method by which to inform a policy

10 judgment on a known or anticipated adverse public welfare effect. For example, a change in an 

11 ecosystem structure and process, such as foliar injury, would be classified as an ecological effect, 

12 with the associated changes in ecosystem services, such as primary productivity, food 

13 availability, forest products, and aesthetics (e.g., scenic viewing), classified as public welfare 

14 effects. Additionally, changes in biodiversity would be classified as an ecological effect, and the 

15 associated changes in ecosystem services—productivity, existence (nonuse) value, recreational 

16 viewing and aesthetics—would also be classified as public welfare effects. 

17 As described in Chapters 4 and 5 of the REA, case study analyses were performed that 

18 link deposition in sensitive ecosystems to changes in a given ecological indicator (e.g., for 

19 aquatic acidification, to changes in acid neutralizing capacity [ANC]) and then to changes in 

20 ecosystems. Appendix 8 of the REA links the changes in ecosystems to the services they

21 provide (e.g., fish species richness and its influence on recreational fishing). To the extent 

22 possible for each targeted effect area, the REA linked ambient concentrations of nitrogen and 

23 sulfur (i.e., ambient air quality indicators) to deposition in sensitive ecosystems (i.e., exposure 
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1 pathways), and then to system response as measured by a given ecological indicator (e.g., lake 

2 and stream acidification as measured by ANC). The ecological effect (e.g., changes in fish 

3 species richness) was then, where possible, associated with changes in ecosystem services and 

4 the corresponding public welfare effects (e.g., recreational fishing). 

5 2. Effects on Ecosystem Services

6 The process used to link ecological indicators to ecosystem services is discussed 

7 extensively in Appendix 8 of the REA. In brief, for each case study area assessed the ecological 

8 indicators are linked to an ecological response that is subsequently linked to associated services 

9 to the extent possible. For example in the case study for aquatic acidification the chosen 

10 ecological indicator is ANC which can be linked to the ecosystem service of recreational fishing. 

11 Although recreational fishing losses are the only service effects that can be independently 

12 quantified or monetized at this time, there are numerous other ecosystem services that may be 

13 related to the ecological effects of acidification.

14 While aquatic acidification is the focus of this proposed standard, the other effect areas 

15 were also analyzed in the REA and these ecosystems are being harmed by nitrogen and sulfur 

16 deposition and will obtain some measure of protection with any decrease in that deposition 

17 regardless of the reason for the decrease. The following summarizes the current levels of specific 

18 ecosystem services for aquatic and terrestrial acidification, and aquatic and terrestrial nutrient 

19 over-enrichment and attempts to quantify and when possible monetize the harm to public 

20 welfare, as represented by ecosystem services, due to nitrogen and sulfur deposition. 

21 a. Aquatic Acidification

22 Acidification of aquatic ecosystems primarily affects the ecosystem services that are 

23 derived from the fish and other aquatic life found in surface waters. In the northeastern United 
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1 States, the surface waters affected by acidification are not a major source of commercially raised 

2 or caught fish; however, they are a source of food for some recreational and subsistence fishers 

3 and for other consumers. Although data and models are available for examining the effects on 

4 recreational fishing, relatively little data are available for measuring the effects on subsistence 

5 and other consumers. Inland waters also provide aesthetic and educational services along with 

6 non-use services, such as existence value (protection and preservation with no expectation of 

7 direct use). In general, inland surface waters such as lakes, rivers, and streams also provide a 

8 number of regulating services, playing a role in hydrological regimes and climate regulation. 

9 There is little evidence that acidification of freshwaters in the northeastern United States has 

10 significantly degraded these specific services; however, freshwater ecosystems also provide 

11 biological control services by providing environments that sustain delicate aquatic food chains. 

12 The toxic effects of acidification on fish and other aquatic life impair these services by disrupting 

13 the trophic structure of surface waters. Although it is difficult to quantify these services and how 

14 they are affected by acidification, it is worth noting that some of these services may be captured 

15 through measures of provisioning and cultural services. For example, these biological control 

16 services may serve as “intermediate” inputs that support the production of “final” recreational 

17 fishing and other cultural services.

18 As summarized in Chapter 4 of the Policy Assessment, recent studies indicate that 

19 acidification of lakes and streams can result in significant loss in economic value. For example, 

20 data indicate that more than 9% of adults in the northeastern part of the country participate 

21 annually in freshwater fishing yielding 140 million freshwater fishing days. Each fishing day

22 has an estimated average value per day of $35. Therefore, the implied total annual value of 

23 freshwater fishing in the northeastern United States was $5 billion in 2006. Embedded in these 
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1 numbers is a degree of harm to recreational fishing services due to acidification that has occurred 

2 over time. These harms have not been quantified on a regional scale; however, a case study was 

3 conducted in the Adirondacks area (US EPA, 2011, section 4.4.2). 

4 In the Adirondacks case study, estimates of changes in recreational fishing services were 

5 determined, as well as changes more broadly in “cultural” ecosystem services (including 

6 recreational, aesthetic, and nonuse services). First, the MAGIC model (US EPA, 2009,

7 Appendix 8 and section 2.2) was applied to 44 lakes to predict what ANC levels would be under 

8 both “business as usual” conditions (i.e., allowing for some decline in deposition due to existing 

9 regulations) and pre-emission (i.e., background) conditions. Second, to estimate the recreational 

10 fishing impacts of aquatic acidification in these lakes, an existing model of recreational fishing 

11 demand and site choice was applied. This model predicts how recreational fishing patterns in the 

12 Adirondacks would differ and how much higher the average annual value of recreational fishing 

13 services would be for New York residents if lake ANC levels corresponded to background 

14 (rather than business as usual) conditions. To estimate impacts on a broader category of cultural 

15 (and some provisioning) ecosystem services, results from the Banzhaf et al (2006) valuation 

16 survey of New York residents were adapted and applied to this context. The survey used a 

17 contingent valuation approach to estimate the average annual household WTP for future 

18 reductions in the percent of Adirondack lakes impaired by acidification. The focus of the survey 

19 was on impacts on aquatic resources. Pretesting of the survey indicated that respondents 

20 nonetheless tended to assume that benefits would occur in the condition of birds and forests as 

21 well as in recreational fishing. 

22 By extrapolating the 44 lake Adiriondack case study to all 3,000 Adirondack lakes and by 

23 applying the willingness to pay survey results to all New York residents, the study estimated 
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1 aggregated benefits between $300 and $800 million annually for the equivalent of improving 

2 lakes in the Adirondacks region to an ANC level of 50 μeq/L. The REA estimated 44% of the 

3 Adirondack lakes currently fall below an ANC of 50 μeq/L. Several states have set goals for 

4 improving the acid status of lakes and streams, generally targeting ANC in the range of 50 to 60 

5 μeq/L, and have engaged in costly activities to decrease acidification. 

6 These results imply significant value to the public in addition to those derived from 

7 recreational fishing services. Note that the results are only applicable to improvements in the 

8 Adirondacks valued by residents of New York. If similar benefits exist in other acid-impacted 

9 areas, benefits for the nation as a whole could be substantial. The analysis provides results on 

10 only a subset of the impacts of acidification on ecosystem services and suggests that the overall 

11 impact on these services is likely to be substantial.

12 b. Terrestrial Acidification

13 Chapter 4.4.3 and 4.4.4 of the Policy Assessment review several economic studies of 

14 areas sensitive to terrestrial acidification. Forests in the northeastern United States provide 

15 several important and valuable provisioning ecosystem services, which are reflected in the 

16 production and sales of tree products. Sugar maples are a particularly important commercial 

17 hardwood tree species in the United States, producing timber and maple syrup that provide 

18 hundreds of millions of dollars in economic value annually. Red spruce is also used in a variety 

19 of wood products and provides up to $100 million in economic value annually. Although the 

20 data do not exist to directly link acidification damages to economic values of lost recreational 

21 ecosystem services in forests, these resources are valuable to the public. A recent study, 

22 reviewed in the Policy Assessment, suggests that the total annual value of recreational off-road 

23 driving was more than $9 billion and the value of hunting and wildlife viewing was more than $4 
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1 billion each in the northeastern United States. EPA is not able to quantify at this time the 

2 specific effects on these values of acid deposition, or of any specific reductions in deposition,

3 relative to the effects of many other factors that may affect them.

4 c. Nutrient Enrichment

5 Chapter 4.4.5 and 4.4.6 of the Policy Assessment summarize economic studies of east 

6 coast estuaries affected by nutrient over-enrichment or eutrophication. Estuaries in the eastern 

7 United States are important for fish and shellfish production. The estuaries are capable of 

8 supporting large stocks of resident commercial species, and they serve as the breeding grounds 

9 and interim habitat for several migratory species. To provide an indication of the magnitude of 

10 provisioning services associated with coastal fisheries, from 2005 to 2007, the average value of 

11 total catch was $1.5 billion per year in 15 East Coast states. Estuaries also provide an important 

12 and substantial variety of cultural ecosystem services, including water-based recreational and 

13 aesthetic services. For example, data indicate that 4.8% of the population in coastal states from 

14 North Carolina to Massachusetts participated in saltwater fishing, with a total of 26 million 

15 saltwater fishing days in 2006. Based on estimates in the Policy Assessment, total recreational 

16 value from these saltwater fishing days was approximately $1.3 billion. Recreational 

17 participation estimates for 1999–2000 showed almost 6 million individuals participated in 

18 motorboating in coastal states from North Carolina to Massachusetts. The aggregate value of 

19 these coastal motorboating outings was $2 billion per year. EPA is not able to quantify at this 

20 time the specific effects on these values of nitrogen deposition, or of any specific reductions in 

21 deposition, relative to the effects of many other factors that may affect them.

22 Terrestrial ecosystems can also suffer from nutrient over-enrichment. Each ecosystem is 

23 different in its composition of species and nutrient requirements. Changes to individual 
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1 ecosystems from changes in nitrogen deposition can be hard to assess economically. Relative 

2 recreational values are often determined by public use information. Chapter 4.4.7 of the Policy 

3 Assessment reviewed studies related to park use in California. Data from California State Parks 

4 indicate that in 2002, 68.7% of adult residents participated in trail hiking for an average of 24.1 

5 days per year. The analyses in the Policy Assessment indicate that the aggregate annual benefit 

6 for California residents from trail hiking in 2007 was $11.59 billion. EPA is not able to quantify 

7 at this time the specific effects on these values of nitrogen deposition, or of any specific 

8 reductions in deposition, relative to the effects of many other factors that may affect them.

9 The Policy Assessment also identified fire regulation as a service that could be affected 

10 by nutrient over-enrichment of the Coastal Sage Scrub and Mixed Conifer Forest ecosystems by

11 encouraging growth of more flammable grasses, increasing fuel loads, and altering the fire cycle. 

12 Over the 5-year period from 2004 to 2008, Southern California experienced, on average, over 

13 4,000 fires per year, burning, on average, over 400,000 acres per year. It is not possible at this 

14 time to quantify the contribution of nitrogen deposition, among many other factors, to increased 

15 fire risk.

16 3. Summary

17 Adversity to public welfare can be understood by looking at how deposition of oxides of 

18 nitrogen and sulfur affect the ecological functions of an ecosystem (see II.A.), and then

19 understanding the ecosystem services that are degraded. The monetized value of the ecosystem 

20 services provided by ecosystems that are sensitive to deposition of oxides of nitrogen and sulfur 

21 are in the billions of dollars each year, though it is not possible to quantify or monetize at this 

22 time the effects on these values of nitrogen and sulfur deposition or of any changes in deposition 

23 that may result from new secondary standards. Many lakes and streams are known to be 
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1 degraded by acidic deposition which affects recreational fishing and tourism. Forest growth is 

2 likely suffering from acidic deposition in sensitive areas affecting red spruce and sugar maple 

3 timber production, sugar maple syrup production, hiking, aesthetic enjoyment and tourism. 

4 Nitrogen deposition contributes significantly to eutrophication in many estuaries affecting fish 

5 production, swimming, boating, aesthetic enjoyment, and tourism. Ecosystem services are

6 likely affected by nutrient enrichment in many natural and scenic terrestrial areas, affecting

7 biodiversity, including habitat for rare and endangered species, fire control, hiking, aesthetic 

8 enjoyment, and tourism. 

9 D. Adequacy of the Current Standards

10 An important issue to be addressed in the current review of the secondary standards for 

11 oxides and nitrogen and sulfur standard is whether, in view of the scientific evidence reflected in 

12 the ISA, additional information on exposure and risk discussed in the REA, and conclusions 

13 drawn from the Policy Assessment, the existing standards provide adequate protection. The 

14 Administrator therefore, has considered the extent to which the current standards are adequate 

15 for the protection of public welfare. Having reached the general conclusion that aquatic and 

16 terrestrial ecosystems can be degraded by deposition of oxides of nitrogen and sulfur, it is then 

17 necessary to first evaluate the appropriateness of the current standards to address the ecological 

18 effects of oxides of nitrogen and sulfur as well as the adequacy of the current secondary

19 standards for oxides of nitrogen and sulfur to provide requisite protection by considering to what 

20 degree risks to sensitive ecosystems would be expected to occur in areas that meet the current 

21 standards. Conclusions regarding the adequacy of the current standards are based on the 

22 available ecological effects, exposure and risk-based evidence. In evaluating the strength of this 

23 information, EPA has taken into account the uncertainties and limitations in the scientific 
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1 evidence. This section addresses the adequacy of the current standards to protect against direct 

2 exposure effects on plants from oxides of nitrogen and sulfur, the appropriateness of the current 

3 structure of the standards to address deposition-related effects of oxides of nitrogen and sulfur on 

4 sensitive ecosystems and finally, the adequacy of such standards to protect against adverse 

5 effects related to the deposition of oxides of nitrogen and sulfur. 

6 1. Adequacy of the Current Standards for Direct Effects

7 The current secondary oxides of nitrogen and sulfur standards are intended to protect 

8 against adverse effects to public welfare. For oxides of nitrogen, the current secondary standard 

was set identical to the primary standard39 , e.g. an annual standard set for NO2 to protect against 

10 adverse effects on vegetation from direct exposure to ambient oxides of nitrogen. For oxides of

11 sulfur, the current secondary standard is a 3-hour standard intended to provide protection for 

12 plants from the direct foliar damage associated with atmospheric concentrations of SO2. It is 

13 appropriate to consider whether the current standards are adequate to protect against the direct 

14 effects on vegetation resulting from ambient NO2 and SO2 which were the basis for the current 

15 secondary standards. The ISA concluded that there was sufficient evidence to infer a causal 

16 relationship between exposure to SO2, NO, NO2 and PAN and injury to vegetation. Additional 

17 research on acute foliar injury has been limited and there is no evidence to suggest foliar injury

18 below the levels of the current secondary standards for oxides of nitrogen and sulfur. There is 

19 sufficient evidence to suggest that the levels of the current standards are likely adequate to 

20 protect against direct phytotoxic effects. 

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

 

3 The current primary NO2 standard has recently been changed to the 3 year average of the 98th percentile of the 

annual distribution of the 1 hour daily maximum of the concentration of NO2. The current secondary standard 

remains as it was set in 1971.
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1 This section addresses two concepts necessary to evaluate the current standards in the 

2 context of deposition related effects. First, appropriateness of the current standards is considered 

3 with regard to indicator, form, level and averaging time. This discussion centers around the 

4 ability of the current standards to evaluate and provide protection against deposition related 

5 effects that vary spatially and temporally. It includes particular emphasis on the indicators and 

6 forms of the current standards and the degree to which they are ecologically relevant with regard 

7 to deposition related effects. Second, this section evaluates the current standards in terms of

8 adequacy of protection. 

9 a. Appropriateness

10 The ISA has established that the major effects of concern for this review of the oxides of 

11 nitrogen and sulfur standards are associated with deposition of N and S caused by atmospheric 

12 concentrations of oxides of nitrogen and sulfur. The current standards are not directed toward 

13 depositional effects, and none of the elements of the current NAAQS – indicator, form, 

14 averaging time, and level – are suited for addressing the effects of N and S deposition. 

15 Five issues arise that call into question the ecological relevance of the structure of the 

16 current secondary standards for oxides of nitrogen and sulfur. 

17 (1) The current SO2 secondary standard (0.5 ppm SO2 over a 3-hour average) does not utilize an 

18 exposure period that is relevant for ecosystem impacts. The majority of deposition related 

19 impacts are associated with depositional loads that occur over periods of months to years. 

20 This differs significantly from exposures associated with hourly concentrations of SO2 as 

21 measured by the current secondary standard. By addressing short-term concentrations, the 

22 current SO2 secondary standard, while protective against direct foliar effects from gaseous 

23 oxides of sulfur, does not take into account the findings of effects in the ISA, which notes the 
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1 relationship between annual deposition of S and acidification effects which are likely to be 

2 more severe and widespread than phytotoxic effects under current ambient conditions, and 

3 include effects from long term deposition as well as short term. Acidification is a process 

4 which occurs over time, as the ability of an aquatic system to counteract acidic inputs is 

5 reduced as natural buffers are used more rapidly than they can be replaced through geologic 

6 weathering. The relevant period of exposure for ecosystems is therefore not the exposures 

7 captured in the short averaging time of the current SO2 secondary standard. The current 

8 secondary standard for oxides of nitrogen is an annual standard (0.053 ppm averaged over 1 

9 year) and as such is more ecologically relevant.

10 (2) Current standards do not utilize appropriate atmospheric indicators. NO2 and SO2 are used 

11 as the component of oxides of nitrogen and sulfur that are measured, but they do not provide 

12 a complete link to the direct effects on ecosystems from deposition of oxides of nitrogen and 

13 sulfur as they do not capture all relevant chemical species of oxidized nitrogen and oxidized 

14 sulfur that contribute to deposition. The ISA provides evidence that deposition related 

15 effects are linked with total nitrogen and total sulfur deposition, and thus all forms of 

16 oxidized nitrogen and oxidized sulfur that are deposited will contribute to effects on 

17 ecosystems. Thus, by using atmospheric NO2 and SO2 concentrations as indicators, the 

18 current standards address only a fraction of total atmospheric oxides of nitrogen and sulfur, 

19 and do not take into account the effects from deposition of total atmospheric oxides of 

20 nitrogen and sulfur. This suggests that more comprehensive atmospheric indicators should 

21 be considered in designing ecologically relevant standards. 

22 (3) Current standards reflect separate assessments of the two individual pollutants, NO2 and SO2, 

23 rather than assessing the joint impacts of deposition to ecosystems. Recognizing the role that 
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1 each pollutant plays in jointly affecting ecosystem indicators, functions, and services is vital 

2 to developing a meaningful standard. The clearest example of this interaction is in 

3 assessment of the impacts of acidifying deposition on aquatic ecosystems. Acidification in 

4 an aquatic ecosystem depends on the total acidifying potential of the deposition of both N 

5 and S from both atmospheric deposition of oxides of nitrogen and sulfur as well as the inputs 

6 from other sources of N and S such as reduced nitrogen and non-atmospheric sources. It is 

7 the joint impact of the two pollutants that determines the ultimate effect on organisms within 

8 the ecosystem, and critical ecosystem functions such as habitat provision and biodiversity. 

9 Standards that are set independently are less able to account for the contribution of the other 

10 pollutant. This suggests that interactions between oxides of nitrogen and oxides of sulfur 

11 should be a critical element of the conceptual framework for ecologically relevant standards. 

12 There are also important interactions between oxides of nitrogen and sulfur and reduced 

13 forms of nitrogen, which also contribute to acidification and nutrient enrichment. It is 

14 important that the structure of the standards address the role of reduced nitrogen in 

15 determining the ecological effects resulting from deposition of atmospheric oxides of 

16 nitrogen and sulfur. Consideration will also have to be given to total loadings as ecosystems 

17 respond to all sources of N and S.

18 (4) Current standards do not take into account variability in ecosystem sensitivity. Ecosystems 

19 are not uniformly distributed either spatially or temporally in their sensitivity to oxides of 

20 nitrogen and sulfur. Therefore, failure to account for the major determinants of variability, 

21 including geological and soil characteristics related to the sensitivity to acidification or 

22 nutrient enrichment as well as atmospheric and landscape characteristics that govern rates of 

23 deposition, may lead to standards that do not provide requisite levels of protection across 
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1 ecosystems. The current structures of the standards do not address the complexities in the 

2 responses of ecosystems to deposition of oxides of nitrogen and sulfur. Ecosystems contain 

3 complex groupings of organisms that respond in various ways to the alterations of soil and 

4 water that result from deposition of nitrogen and sulfur compounds. Different ecosystems 

5 therefore respond in different ways depending on a multitude of factors that control how 

6 deposition is integrated into the system. For example, the same levels of deposition falling 

7 on limestone dominated soils have a very different effect from those falling on shallow 

8 glaciated soils underlain with granite. One system may over time display no obvious 

9 detriment while the other may experience a catastrophic loss in fish communities. This 

10 degree of sensitivity is a function of many atmospheric factors which control rates of 

11 deposition as well as ecological factors which control how an ecosystem responds to that 

12 deposition. The current standards do not take into account spatial and seasonal variations not 

13 only in depositional loadings but also in sensitivity of ecosystems exposed to those loadings. 

14 Based on the discussion summarized above, the Policy Assessment concludes that the current 

15 secondary standards for oxides of nitrogen and oxides of sulfur are not ecologically relevant 

16 in terms of averaging time, form, level or indicator.

17 b. Adequacy of Protection

18 As described in the Policy Assessment, 2005 ambient conditions indicate that the current 

19 SO2 and NO2 secondary standards are not exceeded (US EPA, 2011, Figures 6-1 and 6-2) in 

20 locations where ecological effects have been observed. In many locations, SO2 and NO2 

21 concentrations are substantially below the levels of the secondary standards. This pattern 

22 suggests that levels of deposition and any effects on ecosystems due to deposition of oxides of 

23 nitrogen and sulfur under recent conditions are occurring even though areas meet or are below 
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1 current standards. In addition, these levels based on conclusions in the REA will not decline in 

2 the future to levels below which it is reasonable to anticipate effects.

3 In determining the adequacy of the current secondary standards for oxides of nitrogen 

4 and sulfur the Policy Assessment considered the extent to which ambient deposition contributes 

5 to loadings in ecosystems. Since the last review of the secondary standard for oxides of nitrogen, 

6 a great deal of information on the contribution of atmospheric deposition associated with 

7 ambient oxides of nitrogen has become available. The REA presents a thorough assessment of 

8 the contribution of oxidized nitrogen to nitrogen deposition throughout the U.S., and the relative 

9 contributions of ambient oxidized and reduced forms of nitrogen. The REA concludes that based 

10 on that analysis, ambient oxides of nitrogen are a significant component of atmospheric nitrogen 

11 deposition, even in areas with relatively high rates of deposition of reduced nitrogen. In 

12 addition, atmospheric deposition of oxidized nitrogen contributes significantly to total nitrogen 

13 loadings in nitrogen sensitive ecosystems. 

14 The ISA summarizes the available studies of relative nitrogen contribution and finds that 

15 in much of the U.S., oxides of nitrogen contribute from 50 to 75 percent of total atmospheric 

16 deposition relative to total reactive nitrogen that includes oxidized and reduced nitrogen species 

17 (US EPA, 2008, section 2.8.4). Although the proportion of total nitrogen loadings associated 

18 with atmospheric deposition of nitrogen varies across locations the ISA indicates that 

19 atmospheric N deposition is the main source of new anthropogenic N to most headwater streams, 

20 high elevation lakes, and low-order streams. Atmospheric N deposition contributes to the total N 

21 load in terrestrial, wetland, freshwater, and estuarine ecosystems that receive N through multiple 

22 pathways. In several large estuarine systems, including the Chesapeake Bay, atmospheric 

23 deposition accounts for between 10 and 40 percent of total nitrogen loadings (US EPA, 2008). 
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1 Atmospheric concentrations of oxides of sulfur account for nearly all S deposition in the 

2 US. For the period 2004–2006, mean S deposition in the U.S. was greatest east of the 

3 Mississippi River with the highest deposition amount, 21.3 kg S/ha-yr, in the Ohio River Valley 

4 where most recording stations reported 3 year averages >10 kg S/ha-yr. Numerous other stations 

5 in the East reported S deposition >5 kg S/ha-yr. Total S deposition in the U.S. west of the 100th 

6 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−

7 followed in 

8 decreasing order by a smaller proportion of dry SO2 and a much smaller proportion of deposition 

as dry SO4

2−

9 . 

10 As discussed throughout the REA (US EPA, 2009 and section IIB above), there are 

11 several key areas of risk that are associated with ambient concentrations of oxides of nitrogen 

12 and sulfur. As noted earlier, in previous reviews of the secondary standards for oxides of 

13 nitrogen and sulfur, the standards were designed to protect against direct exposure of plants to 

14 ambient concentrations of the pollutants. A significant shift in understanding of the effects of 

15 oxides of nitrogen and sulfur has occurred since the last reviews, reflecting the large amount of 

16 research that has been conducted on the effects of deposition of nitrogen and sulfur to 

17 ecosystems. The most significant current risks of adverse effects to public welfare are those 

18 related to deposition of oxides of nitrogen and sulfur to both terrestrial and aquatic ecosystems. 

19 These risks fall into two categories, acidification and nutrient enrichment, which were 

20 emphasized in the REA as most relevant to evaluating the adequacy of the existing standards in 

21 protecting public welfare from adverse ecological effects.

22 i. Aquatic acidification
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1 The focus of the REA case studies was on determining whether deposition of sulfur and 

2 oxidized nitrogen in locations where ambient oxides of nitrogen and sulfur were at or below the 

3 current standards was resulting in acidification and related effects, including episodic 

4 acidification and mercury methylation. Based on the case studies conducted for lakes in the 

5 Adirondacks and streams in Shenandoah National Park (case studies are discussed more fully in 

6 section IIB and US EPA, 2009), there is significant risk to acid sensitive aquatic ecosystems at 

7 atmospheric concentrations of oxides of nitrogen and sulfur at or below the current standards. 

8 The REA also supports strongly a relationship between atmospheric deposition of oxides of 

9 nitrogen and sulfur and loss of ANC in sensitive ecosystems and indicates that ANC is an 

10 excellent indicator of aquatic acidification. The REA also concludes that at levels of deposition 

11 associated with oxides of nitrogen and sulfur concentrations at or below the current standards, 

12 ANC levels are expected to be below benchmark values that are associated with significant 

13 losses in fish species richness.

14 Significant portions of the U.S. are acid sensitive, and current deposition levels exceed 

15 those that would allow recovery of the most acid sensitive lakes in the Adirondacks (US EPA, 

16 2008, Executive Summary). In addition, because of past loadings, areas of the Shenandoah are 

17 sensitive to current deposition levels (US EPA, 2008, Executive Summary). Parts of the West 

18 are naturally less sensitive to acidification and subjected to lower deposition (particularly SOx) 

19 levels relative to the eastern United States, and as such, less focus in the ISA is placed on the 

20 adequacy of the existing standards in these areas, with the exception of the mountainous areas of 

21 the West, which experience episodic acidification due to deposition. 

22 In describing the effects of acidification in the two case study areas the REA uses the 

23 approach of describing benchmarks in terms of ANC values. Many locations in sensitive areas of 
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1 the U.S. have ANC levels below benchmark levels for ANC classified as severe, elevated, or 

2 moderate concern (US EPA, 2011, Figure 2-1). The average current ANC levels across 44 lakes 

3 in the Adirondack case study area is 62.1 μeq/L (moderate concern). However, 44 percent of 

4 lakes had deposition levels exceeding the critical load for an ANC of 50 μeq/L, and 28 percent of 

5 lakes had deposition levels exceeding the (higher) critical load for an ANC of 20 μeq/L (US 

6 EPA, 2009, section 4.2.4.2). This information indicates that almost half of the 44 lakes in the 

7 Adirondacks case study area are at an elevated concern level, and almost a third are at a severe 

8 concern level. These levels are associated with greatly diminished fish species diversity, and 

9 losses in the health and reproductive capacity of remaining populations. Based on assessments of 

10 the relationship between number of fish species and ANC level in both the Adirondacks and 

11 Shenandoah areas, the number of fish species is decreased by over half at an ANC level of 20 

12 μeq/L relative to an ANC level at 100 μeq/L (US EPA, 2009, Figure 4.2-1). When extrapolated 

13 to the full population of lakes in the Adirondacks area using weights based on the EMAP 

14 probability survey (US EPA, 2009, section 4.2.6.1), 36 percent of lakes exceeded the critical load 

15 for an ANC of 50 μeq/L and 13 percent of lakes exceeded the critical load for an ANC of 20 

16 μeq/L.

17 Many streams in the Shenandoah case study area also have levels of deposition that are 

18 associated with ANC levels classified as severe, elevated, or moderate concern. The average

19 ANC under recent conditions for the 60 streams evaluated in the Shenandoah case study area is 

20 57.9 μeq/L, indicating moderate concern. However, 85 percent of these streams had recent 

21 deposition exceeding the critical load for an ANC of 50 μeq/L, and 72 percent exceeded the 

22 critical load for an ANC of 20 μeq/L. As with the Adirondacks area, this information suggests 

23 that significant numbers of sensitive streams in the Shenandoah area are at risk of adverse 
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1 impacts on fish populations under recent conditions. Many other streams in the Shenandoah area 

2 are also likely to experience conditions of elevated to severe concern based on the prevalence in 

3 the area of bedrock geology associated with increased sensitivity to acidification suggesting that 

4 effects due to stream acidification could be widespread in the Shenandoah area (US EPA, 2009, 

5 section 4.2.6.2).

6 In addition to these chronic acidification effects, the ISA notes that “consideration of 

7 episodic acidification greatly increases the extent and degree of estimated effects for acidifying 

8 deposition on surface waters.” (US EPA, 2008, section 3.2.1.6) Some studies show that the 

9 number of lakes that could be classified as acid-impacted based on episodic acidification is 2 to 3 

10 times the number of lakes classified as acid-impacted based on chronic ANC. These episodic 

11 acidification events can have long term effects on fish populations (US EPA, 2008, section 

12 3.2.1.6). Under recent conditions, episodic acidification has been observed in locations in the 

13 eastern U.S. and in the mountainous western U.S. (US EPA, 2008, section 3.2.1.6). 

14 The ISA, REA and Policy Assessment all conclude that the current standards are not 

15 adequate to protect against the adverse impacts of aquatic acidification on sensitive ecosystems. 

16 In the ISA it is noted that significant portions of the U.S. are acid sensitive, and that current 

17 deposition levels exceed those that would allow recovery of the most acid sensitive lakes in the 

18 Adirondacks (US EPA, 2008, Executive Summary). In addition, because of past loadings, areas 

19 of the Shenandoah are sensitive to current deposition levels (US EPA, 2008, Executive 

20 Summary). A recent survey, as reported in the ISA, found sensitive streams in many locations in 

21 the U.S., including the Appalachian Mountains, the Coastal Plain, and the Mountainous West 

22 (US EPA, 2008, section 4.2.2.3). In these sensitive areas, between 1 and 6 percent of stream 

23 kilometers are chronically acidified. The REA further concludes that both the Adirondack and 
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1 Shenandoah case study areas are currently receiving deposition from ambient oxides of nitrogen

2 and sulfur in excess of their ability to neutralize such inputs. In addition based on the current 

3 emission scenarios, forecast modeling out to the year 2020 as well as 2050 indicates a large 

4 number of streams in these areas will still be adversely impacted (section IIB). Based on these 

5 considerations, the Policy Assessment concludes that the current secondary NAAQS for oxides 

6 of nitrogen and sulfur do not provide adequate protection of sensitive ecosystems with regard to 

7 aquatic acidification. 

8 ii. Terrestrial acidification

9 Based on the terrestrial acidification case studies (Kane Experimental Forest in 

10 Pennsylvania and Hubbard Brook Experimental Forest described in section IIB) on sugar maple 

11 and red spruce habitat, the REA concludes that there is significant risk to sensitive terrestrial 

12 ecosystems from acidification at atmospheric concentrations of NOx and SOX at or below the 

13 current standards. The ecological indicator selected for terrestrial acidification is the base 

14 cation to aluminum ratio (BC:Al), which has been linked to tree health and growth. The results 

15 of the REA strongly support a relationship between atmospheric deposition of oxides of nitrogen 

16 and sulfur and BC:Al, and that BC:Al is a good indicator of terrestrial acidification. At levels of 

17 deposition associated with oxides of nitrogen and sulfur concentrations at or below the current 

18 standards, BC:Al levels are expected to be below benchmark values that are associated with 

19 significant effects on tree health and growth. Such degradation of terrestrial ecosystems could 

20 affect ecosystem services such as habitat provisioning, endangered species, goods production 

21 (timber, syrup, etc.) among others. 

22 Many locations in sensitive areas of the U.S. have BC:Al levels below benchmark levels 

23 classified as providing low to intermediate levels of protection to tree health. At a BC:Al ratio of 
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1 1.2 (intermediate level of protection), red spruce growth can be reduced by 20 percent. At a 

2 BC:Al ratio of 0.6 (low level of protection), sugar maple growth can be decreased by 20 percent. 

3 The REA did not evaluate broad sensitive regions. However, in the sugar maple case study area 

4 (Kane Experimental Forest), recent deposition levels are associated with a BC:Al ratio below 

5 1.2, indicating between intermediate and low level of protection, which would indicate the 

6 potential for a greater than 20 percent reduction in growth. In the red spruce case study area 

7 (Hubbard Brook Experimental Forest), recent deposition levels are associated with a BC:Al ratio 

8 slightly above 1.2, indicating slightly better than an intermediate level of protection (US EPA, 

9 2009, section 4.3.5.1). 

10 Over the full range of sugar maple, 12 percent of evaluated forest plots exceeded the 

11 critical loads for a BC:Al ratio of 1.2, and 3 percent exceeded the critical load for a BC:Al ratio 

12 of 0.6. However, there was large variability across states. In New Jersey, 67 percent of plots 

13 exceeded the critical load for a BC:Al ratio of 1.2, while in several states on the outskirts of the 

14 range for sugar maple (e.g. Arkansas, Illinois) no plots exceeded the critical load for a BC:Al 

15 ratio of 1.2. For red spruce, overall 5 percent of plots exceeded the critical load for a BC:Al ratio 

16 of 1.2, and 3 percent exceeded the critical load for a BC:Al ratio of 0.6. In the major red spruce 

17 producing states (Maine, New Hampshire, and Vermont), critical loads for a BC:Al ratio of 1.2 

18 were exceeded in 0.5, 38, and 6 percent of plots.

19 The ISA, REA and Policy Assessment all conclude that the current standards are not 

20 adequate to protect against the adverse impacts of terrestrial acidification on sensitive 

21 ecosystems. The main drawback, as stated in the REA and Policy Assessment, with the 

22 understanding of terrestrial acidification lies in the availability of data by which we can predict 

23 critical loads and that the data are based on laboratory responses rather than field measurements. 
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1 Other stressors that are present in the field but that are not present in the laboratory may

2 confound this relationship. The REA does however, conclude that the case study results, when 

3 extended to a 27 state region, show that nitrogen and sulfur acidifying deposition in the sugar 

4 maple and red spruce forest areas caused the calculated Bc/Al ratio to fall below 1.2 (the 

5 intermediate level of protection) in 12% of the sugar maple plots and 5% of the red spruce plots; 

6 however, results from individual states ranged from 0 to 67% of the plots for sugar maple and 0 

7 to 100% of the plots for red spruce fell below the Bc/Al ratio of 1.2.

8 iii. Terrestrial nutrient enrichment

9 Nutrient enrichment effects are due to nitrogen loadings from both atmospheric and non10 atmospheric sources. Evaluation of nutrient enrichment effects requires an understanding that 

11 nutrient inputs are essential to ecosystem health and that specific long term levels of nutrients in 

12 a system affect the types of species that occur over long periods of time. Short term additions of 

13 nutrients can affect species competition, and even small additions of nitrogen in areas that are 

14 traditionally nutrient poor can have significant impacts on productivity as well as species 

15 composition. Most ecosystems in the United States are nitrogen-limited, so regional decreases 

16 in emissions and deposition of airborne nitrogen compounds could lead to some decrease in 

17 growth of the vegetation that surrounds the targeted aquatic system but as discussed below 

18 evidence for this is mixed. Whether these changes in plant growth are seen as beneficial or 

19 adverse will depend on the nature of the ecosystem being assessed. 

20 Information on the effects of changes in nitrogen deposition on forestlands and other 

21 terrestrial ecosystems is very limited. The multiplicity of factors affecting forests, including other 

22 potential stressors such as ozone, and limiting factors such as moisture and other nutrients, 

23 confound assessments of marginal changes in any one stressor or nutrient in forest ecosystems. 
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1 The ISA notes that only a fraction of the deposited nitrogen is taken up by the forests, most of 

2 the nitrogen is retained in the soils (US EPA, 2008, section 3.3.2.1). In addition, the ISA 

3 indicates that forest management practices can significantly affect the nitrogen cycling within a 

4 forest ecosystem, and as such, the response of managed forests to NOx deposition will be 

5 variable depending on the forest management practices employed in a given forest ecosystem 

6 (US EPA, 2008, Annex C C.6.3). Increases in the availability of nitrogen in N-limited forests 

7 via atmospheric deposition could increase forest production over large non-managed areas, but 

8 the evidence is mixed, with some studies showing increased production and other showing little 

9 effect on wood production (US EPA, 2008, section 3.3.9). Because leaching of nitrate can 

10 promote cation losses, which in some cases create nutrient imbalances, slower growth and 

11 lessened disease and freezing tolerances for forest trees, the net effect of increased N on forests 

12 in the U.S. is uncertain (US EPA, 2008, section 3.3.9).

13 The scientific literature has many examples of the deleterious effects caused by excessive 

14 nitrogen loadings to terrestrial systems. Several studies have set benchmark values for levels of 

15 N deposition at which scientifically adverse effects are known to occur. Large areas of the 

16 country appear to be experiencing deposition above these benchmarks. The ISA indicates that 

17 studies have found that at 3.1 kg N/ha/yr, the community of lichens begins to change from 

18 acidophytic to tolerant species; at 5.2 kg N/ha/yr, the typical dominance by acidophytic species 

19 no longer occurs; and at 10.2 kg N/ha/yr, acidophytic lichens are totally lost from the 

20 community. Additional studies in the Colorado Front Range of the Rocky Mountain National 

21 Park support these findings. These three values (3.1, 5.2, and 10.2 kg/ha/yr) are one set of 

22 ecologically meaningful benchmarks for the mixed conifer forest (MCF) of the pacific coast 

23 regions. Nearly all of the known sensitive communities receive total nitrogen deposition levels 
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1 above the 3.1 N kg/ha/yr ecological benchmark according to the12 km, 2002 CMAQ/NADP 

2 data, with the exception of the easternmost Sierra Nevadas. MCFs in the southern portion of the 

3 Sierra Nevada forests and nearly all MCF communities in the San Bernardino forests receive 

4 total nitrogen deposition levels above the 5.2 N kg/ha/yr ecological benchmark. 

5 Coastal Sage Scrub communities (CSS) are also known to be sensitive to community

6 shifts caused by excess nitrogen loadings. Studies have investigated the amount of nitrogen 

7 utilized by healthy and degraded CSS systems. In healthy stands, the authors estimated that 3.3 

8 kg N/ha/yr was used for CSS plant growth. It is assumed that 3.3 kg N/ha/yr is near the point 

9 where nitrogen is no longer limiting in the CSS community and at above which level community

10 changes occur, including dominance by invasive species and loss of coastal sage scrub. 

11 Therefore, this amount can be considered an ecological benchmark for the CSS community. The 

12 majority of the known CSS range is currently receiving deposition in excess of this benchmark. 

13 Thus, the REA concludes that recent conditions where oxides of nitrogen ambient concentrations 

14 are at or below the current oxides of nitrogen secondary standards are not adequate to protect 

15 against anticipated adverse impacts from N nutrient enrichment in sensitive ecosystems. 

16 iv. Aquatic nutrient enrichment

17 The REA aquatic nutrient enrichment case studies focused on coastal estuaries and 

18 revealed that while current ambient loadings of atmospheric oxides of nitrogen are contributing 

19 to the overall depositional loading of coastal estuaries, other non-atmospheric sources are 

20 contributing in far greater amounts in total, although atmospheric contributions are as large as 

21 some other individual source types. The ability of current data and models to characterize the 

22 incremental adverse impacts of nitrogen deposition is limited, both by the available ecological 

23 indicators, and by the inability to attribute specific effects to atmospheric sources of nitrogen. 
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1 The REA case studies used as the ecological indicator for aquatic nutrient enrichment an index 

2 of eutrophication known as the Assessment of Estuarine Trophic Status Eutrophication Index 

3 (ASSETS EI). This index is a six level index characterizing overall eutrophication risk in a 

4 waterbody. This indictor is not sensitive to changes in nitrogen deposition within a single level 

5 of the index. In addition, this type of indicator does not reflect the impact of nitrogen deposition 

6 in conjunction with other sources of nitrogen. 

7 Based on the above considerations, the REA concludes that the ASSETS EI is not an 

8 appropriate ecological indicator for estuarine aquatic eutrophication and that additional analysis 

9 is required to develop an appropriate indicator for determining the appropriate levels of 

10 protection from N nutrient enrichment effects in estuaries related to deposition of oxides of 

11 nitrogen. As a result, EPA is unable to make a determination as to the adequacy of the existing 

12 secondary oxides of nitrogen standard in protecting public welfare from N nutrient enrichment 

13 effects in estuarine aquatic ecosystems.

14 Additionally, nitrogen deposition can alter species composition and cause eutrophication 

15 in freshwater systems. In the Rocky Mountains, for example, deposition loads of 1.5 to 2 

16 kg/ha/yr which are well within current ambient levels are known to cause changes in species 

17 composition in diatom communities indicating impaired water quality (US EPA, 2008, section 

18 3.3.5.3). This suggests that the existing secondary standard for oxides of nitrogen does not 

19 protect such ecosystems and their resulting services from impairment.

20 v. Other effects

21 An important consideration in looking at the effects of deposition of oxides of sulfur in 

22 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-

23
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and mercury, and where mercury is present, increased availability of SO4

2-

1 results in increased 

2 production of MeHg. There is increasing evidence on the relationship between sulfur deposition 

3 and increased methylation of mercury in aquatic environments; this effect occurs only where 

4 other factors are present at levels within a range to allow methylation. The production of 

5 methylmercury requires the presence of sulfate and mercury, but the amount of methylmercury

6 produced varies with oxygen content, temperature, pH, and supply of labile organic carbon (US 

7 EPA, 2008, section 3.4). In watersheds where changes in sulfate deposition did not produce an 

8 effect, one or several of those interacting factors were not in the range required for meaningful 

9 methylation to occur (US EPA, 2008, section 3.4). Watersheds with conditions known to be 

10 conducive to mercury methylation can be found in the northeastern United States and 

11 southeastern Canada (US EPA, 2009, section 6). 

12 With respect to sulfur deposition and mercury methylation, the final ISA determined: The 

13 evidence is sufficient to infer a causal relationship between sulfur deposition and increased 

14 mercury methylation in wetlands and aquatic environments. However, EPA did not conduct a 

15 quantitative assessment of the risks associated with increased mercury methylation under current 

16 conditions. As such, EPA is unable to make a determination as to the adequacy of the existing 

17 SO2 secondary standards in protecting against welfare effects associated with increased mercury

18 methylation.

19 vi. Summary of Adequacy Considerations

20 In summary, the Policy Assessment concludes that currently available scientific evidence 

21 and assessments clearly call into question the adequacy of the current standards with regard to 

22 deposition-related effects on sensitive aquatic and terrestrial ecosystems, including acidification 

23 and nutrient enrichment. Further, the Policy Assessment recognizes that the elements of the 
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1 current standards -- indicator, averaging time, level and form – are not ecologically relevant, and 

2 are thus not appropriate for standards designed to provide such protection. Thus, the Policy

3 Assessment concludes that consideration should be given to establishing a new ecologically

4 relevant multi-pollutant, multimedia standard to provide appropriate protection from deposition5 related ecological effects of oxides of nitrogen and sulfur on sensitive ecosystems, with a focus 

6 on protecting against adverse effects associated with acidifying deposition in sensitive aquatic 

7 ecosystems.

8 3. CASAC Views

9 In a letter to the Administrator (Russell and Samet 2011a), the CASAC Oxides of 

10 Nitrogen and Oxides of Sulfur Panel, with full endorsement of the chartered CASAC, 

11 unanimously concluded that:

12 EPA staff has demonstrated through the Integrated Science Assessment (ISA), 

13 Risk and Exposure Characterization (REA) and the draft PA that ambient NOx 

14 and SOx can have, and are having, adverse environmental impacts. The Panel 

15 views that the current NOx and SOx secondary standards should be retained to 

16 protect against direct adverse impacts to vegetation from exposure to gas phase

17 exposures of these two families of air pollutants. Further, the ISA, REA and draft 

18 PA demonstrate that adverse impacts to aquatic ecosystems are also occurring due 

19 to deposition of NOx and SOx. Those impacts include acidification and 

20 undesirable levels of nutrient enrichment in some aquatic ecosystems. The levels 

21 of the current NOx and SOx secondary NAAQS are not sufficient, nor the forms 

22 of those standards appropriate, to protect against adverse depositional effects; thus 

23 a revised NAAQS is warranted.

24

25 In addition, with regard to the joint consideration of both oxides of nitrogen and oxides 

26 of sulfur as well as the consideration of deposition related effects, CASAC concluded that the 

27 Policy Assessment had developed a credible methodology for considering such effects. The 

28 Panel stated that “the Policy Assessment develops a framework for a multi-pollutant, multimedia 

29 standard that is ecologically relevant and reflects the combined impacts of these two pollutants as 

30 they deposit to sensitive aquatic ecosystems.”
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1 4. Administrator’s Proposed Conclusions Concerning Adequacy of Current Standard

2 Based on the above considerations and taking into account CASAC advice, the 

3 Administrator recognizes that the purpose of the secondary standard is to protect against 

4 “adverse” effects resulting from exposure to oxides of nitrogen and sulfur, discussed above in 

5 section IIA. The Administrator also recognizes the need for conclusions as to the adequacy of the 

6 current standards for both direct and deposition related effects as well as conclusions as to the 

7 appropriateness and ecological relevance of the current standards. 

8 In considering what constitutes an ecological effect that is also adverse to the public 

9 welfare, the Administrator took into account the ISA conclusions regarding the nature and 

10 strength of the effects evidence, the risk and exposure assessment results, the degree to which the 

11 associated uncertainties should be considered in interpreting the results, the conclusions 

12 presented in the Policy Assessment, and the views of CASAC and members of the public. On 

13 these bases, the Administrator concludes that the current secondary standards are adequate to 

14 protect against direct phytotoxic effects on vegetation. Thus, the Administrator proposes to 

retain the current secondary standard for oxides of nitrogen at 53 ppb,415 annual average 

16 concentration, measured in the ambient air as NO2, and the current secondary standard for oxides 

17 of sulfur at 0.5 ppm, 3-hour average concentration, measured in the ambient air as SO2. 

18 With regard to deposition-related effects, the Administrator has first to consider the 

19 appropriateness of the structure of the current standards to address ecological effects of concern. 

20 Based on the evidence as well as considering the advice given by CASAC on this matter, the 

21 Administrator concludes that the elements of the current standards are not ecologically relevant 

22 and thus are not appropriate to provide protection of ecosystems. On the subject of adequacy of 

 

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).
Interagency Working Comments on Draft Language under EO12866 and 

13563 Interagency Review. Subject to Further Policy Review.

102

1 protection with regard to deposition-related effects, the Administrator considered the full nature 

2 of ecological effects related to the deposition of ambient oxides of nitrogen and sulfur into 

3 sensitive ecosystems across the country. Her conclusions are based on the evidence presented in 

4 the ISA with regard to acidification and nutrient enrichment effects, the findings of the REA with 

5 regard to scope and severity of the current and likely future effects of deposition, the synthesis of 

6 both the scientific evidence and risk and exposure results in the Policy Assessment as to the 

7 adequacy of the current standards, and the advice of both CASAC and the public. After such

8 consideration, the Administrator concludes that current levels of oxides of nitrogen and sulfur are 

9 sufficient to cause acidification of both aquatic and terrestrial ecosystems, nutrient enrichment of 

10 terrestrial ecosystems and contribute to nutrient enrichment effects in estuaries that could be 

11 considered adverse, and the current secondary standards do not provide adequate protection from 

12 such effects. 

13 Having reached these conclusions, the Administrator determines that it is appropriate to 

14 consider alternative standards that are ecologically relevant. These considerations support the 

15 conclusion that the current secondary standards is neither appropriate nor adequate to protect 

16 against deposition related effects. The Administrator’s consideration of such alternative 

17 standards is discussed below in Section III.