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1 ENVIRONMENTAL PROTECTION AGENCY 1 40 CFR Part 50 2 [EPA - HQ - OAR- 2007 - 1145] 3 RIN: 2060 - AO72 4 Secondary National Ambient Air Quality Standards for Oxides of Nitrogen and Sulfur 5 6 AGENCY: Environmental Protection Agency (EPA). 7 ACTION: Proposed rule. 8 SUMMARY: Based on its review of the air quality criteria and the secondary national ambient 9 air quality

standards (NAAQS) for oxides of nitrogen and oxides of sulfur, EPA proposes to 10 retain the current nitrogen dioxide (NO 2 ) and sulfur dioxide (SO 2 ) secondary stand ards to 11 provide requisite protection for the direct effects on vegetation resulting from exposure to 12 gaseous oxides of nitrogen and sulfur in the ambient air. With regard to protection from the 13 deposition of oxides of nitrogen and sulfur to sensitive aquatic and terrestrial ecosy stems, 14 including acidification and nutrient enrichment effects, EPA is deferring the setting of a new 15 multi

- pollutant secondary standard in this review, and is proposing not to set such a secondary 16 standard at this time

. Instead, EPA has decided to undertake a field pilot program to gather and 17 anal yz e additional relevant data so as to enhance the Agency ’s understanding of the degree of 18 protectiveness that a new multi

- pollutant standard, defined in terms of an aquatic acidification 19 index (AAI), would afford and to support development of an appropriate monitoring network for 20 such a standard. The EPA solicits comment on the framework of such a standard and on the 21 design of the field pilot program. 22Draft

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2 DATES: Written comments on this prop osed rule must be received b y [insert date 60 days 1 after date of publication in the Federal Register]
. 2 Public Hearings : The EPA intends to hold a public hearing around the end of August to 3 earl y September and will announce in a separate Federal Register n otice the date. time, and 4 address of the public hearing on this proposed rule. 5 ADDRESSES: Submit your comments, identified by Docket ID No. EPA

- HQ
- OAR
- 2007
- 6 1145, by

one of the following methods: 7

 www.regulations.gov : Follow the on

- line instructions for submitting comments. 8

 Email: a - and

- r - Docket@epa.gov

. 9

 Fax: 202
- 566
- 1741. 10

 Mail: Docket No. EPA

- HQ
- OAR
- 2007
- 1145, Environmental Protection Agency , Mail 11 code 6102T, 1200 Penns yl vania Ave., NW., Washington, DC 20460. Please include a 12 total of two copies. 13

 Hand Delivery : Docket No. EPA

- HQ
- OAR
- 2007
- 1145, Environmental Protection 14 Agency , EPA West, Room 3334, 1301 Constitution Ave., NW, Washington, DC. Such 15 deliveries are onl y ac cepted during the Docket’s normal hours of operation, and special 16 arrangements should be made for deliveries of boxed information. 17 Instructions: Direct your comments to Docket ID No. EPA

- HQ
- OAR
- 2007
- 1145. The 18 EPA’s policy

is that all comments received wi ll be included in the public docket without change 19 and may

be made available online at www.regulations.gov , including an y personal information 20 provided, unless the comment includes information claimed to be Confidential Business 21 In formation (CBI) or other information whose disclosure is restricted b y statute. Do not submit 22 information that you consider to be CBI or otherwise protected through www.regulations.gov or 23Draft

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3 email. Th e www.regulations.gov website is an “anony mous access” sy stem, which means EPA 1 will not know your identity or contact information unless you provide it in the body of your 2 comment. If you send an email comment dire ctl y to EPA without going through 3 www.regulations.gov , your email address will be automatically captured and included as part of 4 the comment that is placed in the public docket and made available on the Internet. I f you 5 submit an electronic comment, EPA recommends that you include your name and other contact 6 information in the body of your comment and with any disk or CD

- ROM you submit. I f EPA 7 cannot read your comment due to technical difficulties and cannot contac t you for clarification, 8 EPA may

not be able to consider your comment. Electronic files should avoid the use of special 9 characters, an y form of encry ption, and be free of any

defects or viruses. For additional 10 information about EPA’s public docket, visit the EPA Docket Center homepage at 11 http://www.epa.gov/epahome/dockets.htm

. 12 Docket : All documents in the docket are listed in the www.regulations.gov index. 13 Al though listed in the index, some information is not publicly available, e.g., CBI or other 14 information whose disclosure is restricted b y statute. Certain other material, such as cop yri ghted 15 material, will be publicly available onl y in hard copy . Publicly available docket materials are 16 available either electronically in www.regulations.gov or in hard cop y at the Air and Radiation 17 Docket and Information Center, EPA/DC, EPA West, Room 3334, 1301 Constitution Ave., NW, 18 Washington, DC. The Public Reading Room is open from 8:30 a.m. to 4:30 p.m., Monday 19 through Friday , excluding legal holiday s. The telephone number for the Public Reading Room is 20 (202) 566

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

- 1742. 22Draft

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4 FOR FURTHER INFORMATION CONTACT: Dr. Richard Scheffe, Office of Air Quality 1 Planning and Standards, U.S. Environmental Protection Agency , Mail code C304

- 02, Research 2 Triangle Park, NC 27711; telephone: 919

- 541
- 46 50; fax: 919

- 541
- 2357; email: 3 scheffe.rich@epa.gov. 4 SUPPLEMENTARY INFORMATION: 5 General Information 6 What Should IConsider as I Prepare M y Comments for EPA? 7 1. Submitting CBI. Do not submit this information to EPA through www.regulations.gov or 8 email. Clearl y mark the part or all of the information that you claim to be CBI. For CBI 9 information in a disk or CD ROM that you mail to EPA, mark the outside of the disk or 10 CD ROM as CBI and then identify electronically

within the disk or CD ROM the specific 11 information that is claimed as CBI. In addition to one complete version of the comment 12 that includes information claimed as CBI, a cop y of the comment that does not contain 13 the informati on claimed as CBI must be submitted for inclusion in the public docket. 14 In formation so marked will not be disclosed except in accordance with procedures set 15 forth in 40 CFR part 2. 16 2. Tips for Preparing Your Comments . When submitting comments, remember to: 17

 Id entify

the rulemaking by

docket number and other identifying information (subject 18 heading, Federal Register date and page number). 19

 Follow directions

– The Agency may ask you to respond to specific questions or 20 organize comments by referencing a Code of Federal Regulations (CFR) part or section 21 number. 22Draft

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5  Explain why

you agree or disagree, suggest alternatives, and substitute language for your 1 requested changes. 2

 Describe an y assumptions and provide an y technical information and/or data that you 3 used. 4

 If you estimate potential costs or burdens, explain how you arrived at your estimate in 5 sufficient detail to allow for it to be reproduced. 6

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

 Explain your views as clearl y as possible, avoiding the use of profanity or personal 8 threats. 9

 Make sure to submit y our comments by the comment period deadline identified. 10 Availability

of Related Information 11 A number of documents relevant to this rulemaking are available on EPA web sites. The 12 Int egrated Science Assessment for Oxides of Nitrogen and Sulfur

- Ecological Criteria: FINAL 13 REPORT (I SA) is available on EPAs National Center for Environmental Assessment web site. 14 To obtain this document, go to http://www.epa.gov/ncea , and click on Air Quality then click on 15 Oxides of Nitrogen and Sulfur. The Policy Assessment, Risk and Exposure Assessment (REA) , 16 and other related technical documents are available on EPA’s Office of Air Quality Planning and 17 Standards (O AQPS) Technology Transfer Network (TTN) web site. The Policy Assessment is 18 available at http://www.epa.gov/ttn/naaqs/standards/no2so2sec/cr_pa.html, and the exposure and 19 risk assessments and other related technical documents are available at 20 http://www.epa.gov/ttn/naaqs/standards/no2so2sec/cr_rea.html

. These and other related 21 documents are also available for inspection and copying in the EPA docket identified above. 22Draft

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6 Table of Conten ts 1 The following topics are discussed in this preamble: 2 I. Background 3 A. Legislative Requirements 4 B. History of Reviews of NAAQS for Nitrogen Oxides and Sulfur Oxides 5 1. NAAQS for Oxides of Nitrogen 6 2. NAAQS for Oxides of Sulfur 7 C. History of Related Assessments and Agency Actions 8 D. History of the Current Review 9 E. Scope of the Current Review 10 II. Rationale for Proposed Decision on the Adequacy of the Current Secondary Standards 11 A. Ecological Effects 12 1. Effects Associated with Gas

- Phase Oxides of Nitrogen and Sulfur 13 a. Nature of ecos ys tem responses to gas - phase nitrogen and sulfur 14 b. Magnitude of ecos ys tem response to gas - phase nitrogen and sulfur 15 2. Acidification Effects Associated with Deposition of Oxides of N itrogen 16 and Sulfur 17 a. Nature of Acidification

–related Ecosy stem Responses 18 i. Aquatic ecos ys tems 19 ii. Terrestrial ecos ys tems 20 iii. Ecosy stem sensitivity 21 b. Magnitude of Acidification

- related Ecos ys tem Responses 22 i. Aquatic acidification 23 ii. Terrestrial acidification 24 c. Key Uncertainties Associated with Acidification 25 i. Aquatic acidification 26 ii. Terrestrial acidification 27 3. Nutrient Enrichment Effects Associated with Deposition of Oxides of 28 Nitrogen 29 a. Nature of Nutrient Enrichment

- related Ecos ys tem Responses 30 i. Aquatic ecos ys tems 31 ii. Terrestrial ecos ys tems 32 iii. Ecosy stem sensitiv ity to nutrient enrichment 33 b. Magnitude of Nutrient Enrichment

- related Ecos ys tem Responses 34 i. Aquatic ecos ys tems 35 ii. Terrestrial ecos ys tems 36 c. Key Uncertainties Associated with Nutrient Enrichment 37 i. Aquatic ecos ys tems 38 ii. Terrestrial ecos ys tems 39 4. Other Ecological Effec ts 40 B. Risk and Exposure Assessment 41 1. Overview of Risk and Exposure Assessment 42 2. Key Findings 43 a. Air quality anal ys es 44Draft

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7 b. Deposition

- related aquatic acidification 1 c. Deposition

- related terrestrial acidification 2 d. Deposition

- related aquatic nutrient enrichment 3 e. Deposition

- related terrestrial nutrient enrichment 4 f. Additional effects 5 3. Conclusions on Effects 6 C. Adversity of Effects to Public Welfare 7 1. Ecosy stem Services 8 2. Effects on Ecosy stem Services 9 a. Aquatic Acidification 10 b. Terrestrial Acidification 11 c. Nutrient Enrichment 12 3. Summary 13 D. Adequacy of the Current Standards 14 1. Adequacy of the Current Standards for Direct Effects 15 2. Appropriateness and Adequacy of the Current Standards for Deposition

16 related Effects 17 a. Appropriateness 18 b. Adequacy of Protection 19 i. Aquatic Acidification 20 ii. Terrestrial aci dification 21 iii. Terrestrial nutrient e nrichment 22 iv. Aquatic nutrient e nrichment 23 v. Other effects 24 3. CASAC Views 25 4. Administrator’s Proposed Conclusions Concerning Adequacy of Current 26 Standard 27 III. Rationale for Proposed Decision on Alternative Secondary Standard s for Aquatic 28 Acidification 29 A. Ambient Air Indicators 30 1. Oxides of Sulfur 31 2. Oxides of Nitrogen 32 B. Form 33 1. Ecological Indicator 34 2. Linking ANC to Deposition 35 3. Linking Deposition to Ambient Air Indicators 36 4. Aquatic Acidification Index 37 5. Spatial Aggregation 38 a. Ecoregion Sensitivity 39 b. Representative Ecoregion

- specific Factors 40 i. Factor F1 41 a) Acid

- sensitive Ecoregions 42 b) Non
- acid sensitive Ecoregions 43 ii. Factor F2 44 iii. Factors F3 and F4 45 c. Factors in Data

- limited Ecoregions 46Draft

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8 d. Application to Hawaii, Alaska, and the U.S. Territories 1 6. Summary of the AAI Form 2 C. Averaging Time 3 D. Level 4 1. Association Between pH Levels and Target ANC Levels 5 2. ANC Levels Related to Effects on Aquatic Ecos ys tems 6 3. Consideration of Episodic Acidity 7 4. Consideration of Ecos ys tem Response Time 8 5. Prior Examples of Target ANC Levels 9 6. Consideration of Public Welfare Benefits 10 7. Summary of Alternative Levels 11 E. Combined Alternative Levels and Forms 12 F. Characterization of Uncertainties 13 1. Overview of Uncertaint y 14 2. Uncertainties Associated with Data Gaps 15 3. Uncertainties in Modeled Processes 16 G. CASAC Advice 17 H. Administrator’s Proposed Conclusi ons 18 IV
. Field Pilot Program and Ambient Monitoring 19 A. Field Pilot Program 20 a. Objectives 21 b. Overview of Field Pilot Program 22 c. Complementary Measurements 23 d. Complementary Research Efforts 24 e. Implementation Challenges 25 f. Final Monitoring Plan Development and Stakeholder Participation 26 B. Evaluation of Monitoring Methods 27 1. Potential FRMs for SO 2 and p

- SO 4 28 2. Potential FRM for NO y 29 V. Statutory

and Executive Order Reviews 30 A. Executive Order 12866: Regulatory Planning and Review 31 B. Paperwork Reduction Act 32 C. Regulatory Flexibility

Act 33 D. Unfunded Mandates Reform Act 34 E. Executive Order 13132: Federalism 35 F. Executive Order 13175: Consultation and Coordination with Indian Tribal 36 Governments 37 G. Executive Order 13045: Protection of Children From Environmental Health and 38 Safety Risks 39 H. Executive Order 13211: Ac tions that Significantly Affect Energy

Suppl y, 40 Distribution, or Use 41 I. National Technology Transfer and Advancement Act 42 J. Executive Order 12898: Federal Actions to Address Environmental Justice in 43 Minority Populations and L ow - Income Populations 44 References 45 46Draft

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9 I. Background 1 A. Legislative Requirements 2 Two sections of the Clean Air Act (CAA) govern the establishment and revision of the 3 NAAQS. Section 108 (42 U.S.C. section 7408) directs the Administrator to identify and list 4 certain air pollutants and then to issue air quality criteria for those pollutants. The Administrator 5 is to list those air pollutants that in her “judgment, cause or contribute to air pollution which may 6 reasonabl y be anticipated to endanger public health or welfare;” “the presence of whic h in the 7 ambient air results from numerous or diverse mobile or stationary sources;” and “for which . . . 8 [the Administrator] plans to issue air quality criteria...” Air qualit y criteria are intended to 9 “accuratel y reflect the latest scientific knowledge us eful in indicating the kind and extent of all 10 identifiable effects on public health or welfare which may be expected from the presence of [a] 11 pollutant in the ambient air . . .” 42 U.S.C. § 7408(b). Section 109 (42 U.S.C. 7409) directs the 12 Administrator to propose and promulgate “primary” and “secondary” NAAQS for pollutants for 13 which air qualit y criteria are issued. Section 109(b)(1) defines a primary standard as one “the 14 attainment and maintenance of which in the judgment of the Administrator, based on such 15 criteria and allowing an adequate margin of safety , are requisite to protect the public health.” 1 A 16 secondary

standard, as defined in section 109(b)(2), must “specify a level of air qualit y the 17 attainment and maintenance of which, in the judgment of the Administrator, based on such 18 criteria, is requisite to protect the public welfare from any known or anticipated adverse effects 19 associated with the presence of [the] pollutant in the ambient air.” Welfare effects as defined in 20 section 302(h) (42 U.S.C . § 7602(h)) include, but are not limited to, “effects on soils, water, 21

1 The legislative history of section 109 indicates that a primary standard is to be set at “the maximum permissible am bient 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 representativ e 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).Draft

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10 crops, vegetation, man

- made materials, animals, wildlife, weather, visibility and climate, damage 1 to and deterioration of property , and hazards to transportation, as well as effects on economic 2 values and on personal comfort and well

- being.” 3 In setting standards that are “requisite” to protect public health and welfare, as provided 4 in section 109(b), EPA’s task is to establish standards that are neither more nor less stringent 5 than nec essary

for these purposes. In so doing, EPA may not consider the costs of implementing 6 the standards. See generally , Whitman v. American Trucking Associations, 531 U.S. 457, 465

- 7 472, 475
- 76 (2001). Likewise, “[a]ttainability and technological feasibility are not relevant 8 considerations in the promulgation of national ambient air quality standards.” American 9 Petroleum Institute v. Costle, 665 F. 2d at 1185. Section 109(d)(1) requires that “not later than 10 December 31, 1980, and at 5

- ye ar intervals thereaf ter, the Administrator shall complete a 11 thorough review of the criteria published under section 108 and the national ambient air quality 12 standards . . . and shall make such revisions in such criteria and standards and promulgate such 13 new standards as may b e appropriate . . . .” Section 109(d)(2) requires that an independent 14 scientific review committee “shall complete a review of the criteria . . . and the national primary 15 and secondary

ambient air quality

standards . . . and shall recommend to the Administrator any 16 new . . . standards and revisions of existing criteria and standards as may be appropriate . . . .” 17 Since the earl y 1980's, this independent review function has been performed by the Clean Air 18 Scientific Advisory Committee (CASAC). 19 B. History of Reviews of NAAQS for Nitrogen Oxides and Sulfur Oxides 20 1. NAAQS for Oxides of Nitrogen 21 After reviewing the relevant science on the public health and welfare effects associated 22 with oxides of nitrogen, EPA promulgated identical primary and secondary NAAQS for NO2 in 23Draft

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11 April 1971. These standards were set at a level of 0.053 parts per million (ppm) as an annual 1 average (36 FR 8186). In 1982, EPA published Air Quality Criteria for Oxides of Nitrogen (US 2 EPA, 1982), which updated the scientific criteria upon which the initial standards were based. In 3 February

1984 EPA proposed to retain these standards (49 FR 6866). After taking into account 4 public comments, EPA published the final decision to retain these standards in June 1985 (50 FR 5 25532). 6 The EPA began the most recent previous review of the oxides of nitrogen secondary 7 standards in 1987. In November 1991 EPA released an updated draft air quality criteria 8 document ( AQCD ) for CASAC and public review and comment (56 FR 59285), which provided 9 a comprehensi ve assessment of the available scientific and technical information on health and 10 welfare effects associated with NO2 and other oxides of nitrogen. The CASAC reviewed the 11 draft document at a meeting held on July 1, 1993 and concluded in a closure letter t o the 12 Administrator that the document “provides a scientifically balanced and defensible summary of 13 current knowledge of the effects of this pollutant and provides an adequate basis for EPA to 14 make a decision as to the appropriate NAAQS for NO2” (Wolff, 19 93). The Air Qualit y Criteria 15 for Oxides of Nitrogen was then finalized (US EPA, 199 5 a). EPA’s OAQPS also prepared a 16 Staff Paper that summarized and integrated the key studies and scientific evidence contained in 17 the revised AQCD for oxides of nitrogen and identified the critical elements to be considered in 18 the review of the NO2 NAAQS. CASAC reviewed two drafts of the Staff Paper and concluded 19 in a closure letter to the Administrator that the document provided a “scientifically adequate 20 basis for regulatory decisions on nitrogen dioxide” (Wolff, 1995). 21 In October 1995 the Administrator announced her proposed decision not to revise either 22 the primary

or secondary NAAQS for NO2 (60 FR 52874; October 11, 1995). A y ear later, the 23Draft

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12 Administrator made a final determination not to revise the NAAQS for NO2 after careful 1 evaluation of the comments received on the proposal (61 FR 52852; October 8, 1996). While the 2 primary

NO2 standard was revised in January 2010 by supplementing the existing annual 3 standard with the establishment of a new 1

- hour standard (75 FR 6474), the secondary NAAQS 4 for NO2 remains 0.053 ppm (100 micrograms per cubic meter [μg/m3] of air), annual arithmetic 5 average, calculated as the arithmetic mean of the 1

- hour NO2 concentrations. 6 2. NAAQS for Oxides of Sulfur 7 EPA promulgated primary and secondary

NAAQS for SO2 in April 1971 (36 FR 8186). 8 The secondary

standards included a standard set at 0.02 ppm, annual arithmetic mean, and a 3

- 9 hour average standard set at 0.5 ppm, not to be exceeded mor e than once per y ear. These 10 secondary

standards were established solel y on the basis of evidence of adverse effects on 11 vegetation. In 1973, revisions made to Chapter 5 (“Effects of Sulfur Oxide in the Atmosphere on 12 Vegetation”) of Air Quality Criteria fo r Sulfur Oxides (US EPA, 1973) indicated that it could not 13 properl y be concluded that the vegetation injury reported resulted from the average SO2 14 exposure over the growing season, rather than from short

- term peak concentrations. Therefore, 15 EPA proposed ( 38 FR 11355) and then finalized (38 FR 25678) a revocation of the annual mean 16 secondary

standard. At that time, EPA was aware that then

- current concentrations of oxides of 17 sulfur in the ambient air had other public welfare effects, including effects on materials, 18 visibility , soils, and water. However, the available data were considered insufficient to establish 19 a quantitative relationship between specific ambient concentrations of oxides of sulfur and such 20 public welfare effects (38 FR 25679). 21 In 1979, EP A announced that it was revising the AQCD for oxides of sulfur concurrentl y 22 with that for particulate matter (PM) and would produce a combined PM and oxides of sulfur 23Draft

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13 criteria document. Following its review of a draft revised criteria document in August 1 980, 1 CASAC concluded that acid deposition was a topic of extreme scientific complexity because of 2 the difficulty

in establishing firm quantitative relationships among (1) emissions of relevant 3 pollutants (e.g., SO2 and oxides of nitrogen), (2) formation of acidic wet and dry deposition 4 products, and (3) effects on terrestrial and aquatic ecos ys tems. CASAC also noted that acid 5 deposition involves, at a minimum, several different criteria pollutants: oxides of sulfur, oxides 6 of nitrogen, and the fine partic ulate fraction of suspended particles. CASAC felt that any 7 document on this subject should address both wet and dry deposition, since dry

deposition was 8 believed to account for a substantial portion of the total acid deposition problem. 9 For these reasons , CASAC recommended that a separate, comprehensive document on 10 acid deposition be prepared prior to an y consideration of using the NAAQS as a regulatory 11 mechanism for the control of acid deposition. CASAC also suggested that a discussion of acid 12 depositio n be included in the AQCDs for oxides of nitrogen and PM and oxides of sulfur. 13 Following CASAC closure on the AQCD for oxides of sulfur in December 1981, EPA’s OAQPS 14 published a Staff Paper in November 1982, although the paper did not directly assess the issue of 15 acid deposition. Instead, EPA subsequently prepared the following documents to address acid 16 deposition: The Acidic Deposition Phenomenon and Its Effects: Critical Assessment Review 17 Papers, Volumes I and II (US EPA, 1984a, b) and The Acidic Depos ition Phenomenon and Its 18 Effects: Critical Assessment Document (US EPA, 1985) (53 FR 14935

- 14936). These 19 documents, though they were not considered criteria documents and did not undergo CASAC 20 review, represented the most comprehensive summary of scienti fic information relevant to acid 21 deposition completed by EPA at that point. 22Draft

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14 In April 1988 (53 FR 14926), EPA proposed not to revise the existing primary and 1 secondary

standards for SO 2 . This proposed decision with regard to the secondary SO 2 NAAQS 2 was du e to the Administrator’s conclusions that (1) based upon the then

- current scientific 3 understanding of the acid deposition problem, it would be premature and unwise to prescribe any 4 regulatory

control program at that time and (2) when the fundamental scient ific uncertainties had 5 been decreased through ongoing research efforts, EPA would draft and support an appropriate set 6 of control measures. Although EPA revised the primary SO 2 standard in June 2010 by 7 establishing a new 1

- hour standard and revoking the e xisting 24

- hour and annual standards (75 8 FR 35520), no further decisions on the secondary SO 2 standard have been published. 9 C. History of Related Assessments and Agency Actions 10 In 1980, the Congress created the National Acid Precipitation Assessment Pro gram 11 (NAPAP) in response to growing concern about acidic deposition. The NAPAP was given a 12 broad 10

- ye ar mandate to examine the causes and effects of acidic deposition and to explore 13 alternative control options to alleviate acidic deposition and its effec ts. During the course of the 14 program, the NAPAP issued a series of publicl y available interim reports prior to the completion 15 of a final report in 1990 (NAPAP, 1990). 16 In spite of the complexities and significant remaining uncertainties associated with th e 17 acid deposition problem, it soon became clear that a program to address acid deposition was 18 needed. The Clean Air Act Amendments of 1990 included numerous separate provisions related 19 to the acid deposition problem. The primary and most important of the provisions, the 20 amendments to Title IV of the Act, established the Acid Rain Program to reduce emissions of 21 SO 2 by

10 million tons and emissions of nitrogen oxides by 2 million tons from 1980 emission 22 levels in order to achieve reductions over broad geogr aphic regions. In this provision, Congress 23Draft

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15 included a statement of findings that led them to take action, concluding that (1) the presence of 1 acid compounds and their precursors in the atmosphere and in deposition from the atmosphere 2 represents a threat t o natural resources, ecos ys tems, materials, visibility , and public health; (2) 3 the problem of acid deposition is of national and international significance; and (3) current and 4 future generations of Americans will be adversely affected by

delaying measures to remed y the 5 problem. 6 Second, Congress authorized the continuation of the NAPAP in order to assure that the 7 research and monitoring efforts already undertaken would continue to be coordinated and would 8 provide the basis for an impartial assessment of the effectiveness of the Title IV program. 9 Third, Congress considered that further action might be necessary in the long term to 10 address an y problems remaining after implementation of the Title IV program and, reserving 11 judgment on the form that action cou ld take, included Section 404 of the 1990 Amendments 12 (Clean Air Act Amendments of 1990, Pub. L . 101
- 549, § 404) requiring EPA to conduct a study 13 on the feasibility

and effectiveness of an acid deposition standard or standards to protect 14 “sensitive and crit ically sensitive aquatic and terrestrial resources.” At the conclusion of the 15 study , EPA was to submit a report to Congress. Five y ears later, EPA submitted its report, 16 entitled Acid Deposition Standard Feasibility Study : Report to Congress (US EPA, 1995 b ) in 17 fulfillment of this requirement. That Report concluded that establishing acid deposition 18 standards for sulfur and nitrogen deposition may at some point in the future be technically 19 feasible, although appropriate deposition loads for these acidifying chemicals could not be 20 defined with reasonable certaint y at that time. 21 Fourth, the 1990 Amendments also added new language to sections of the CAA 22 pertaining to the scope and application of the secondary NAAQS designed to protect the public 23Draft

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16 welfare. Spe cifically , the definition of “effects on welfare” in Section 302(h) was expanded to 1 state that the welfare effects include effects “...whether caused by transformation, conversion, or 2 combination with other air pollutants.” 3 In 1999, seven Northeastern stat es cited this amended language in Section 302(h) in a 4 petition asking EPA to use its authority under the NAAQS program to promulgate secondary 5 NAAQS for the criteria pollutants associated with the formation of acid rain. The petition stated 6 that this language “clearly references the transformation of pollutants resulting in the inevitable 7 formation of sulfate and nitrate aerosols and/or their ultimate environmental impacts as wet and 8 dry deposition, clearl y signaling Congressional intent that the welfare d amage occasioned b y 9 sulfur and nitrogen oxides be addressed through the secondary standard provisions of Section 10 109 of the Act.” The petition further stated that “recent federal studies, including the NAPAP 11 Biennial Report to Congress: An Integrated Asse ssment, document the continued and increasing 12 damage being inflicted by acid deposition to the lakes and forests of New York, New England 13 and other parts of our nation, demonstrating that the Title IV program had proven insufficient.” 14 The petition also lis ted other adverse welfare effects associated with the transformation of these 15 criteria pollutants, including impaired visibility , eutrophication of coastal estuaries, global 16 warming, and tropospheric ozone and stratospheric ozone depletion. 17 In a related m atter, the Office of the Secretary of the U.S. Department of Interior (DOI) 18 requested in 2000 that EPA initiate a rulemaking proceeding to enhance the air qualit y in 19 national parks and wilderness areas in order to protect resources and values that are bein g 20 adversel y affected b y air pollution. Included among the effects of concern identified in the 21 request were the acidification of streams, surface waters, and/or soils; eutrophication of coastal 22 waters; visibility

impairment; and foliar injury from ozone. 23Draft

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17 In a Federal Register notice in 2001, EPA announced receipt of these requests and asked 1 for comment on the issues raised in them. EPA stated that it would consider an y relevant 2 comments and information submitted, along with the information provided by the petitioners and 3 DOI, before making an y decision concerning a response to these requests for rulemaking (65 FR 4 48699). 5 The 2005 NAPAP report states that “... scientific studies indicate that the emission 6 reductions achieved b y Title IV are not sufficient to allow recovery of acid

- sensitive ecos ys tems. 7 Estimates from the literature of the scope of additional emission reductions that are necessary in 8 order to protect acid

- sensitive ecosy stems range from approximately 40
- 80% beyond full 9 implementation of Title IV....” The results of the modeling presented in this Report to Congress 10 indicate that broader recovery is not predicted without additional emission reductions” (NAPAP , 11 2005 ). 12 Given the state of the science as described in the ISA, REA, and in other rec ent reports, 13 such as the NAPAP reports noted above, EPA has decided, in the context of evaluating the 14 adequacy

of the current NO2 and SO2 secondary standards in this review, to revisit the question 15 of the appropriateness of setting secondary NAAQS to addre ss remaining known or anticipated 16 adverse public welfare effects resulting from the acidic and nutrient deposition of these criteria 17 pollutants. 18 D. History of the Current Review 19 The EPA initiated this current review in December 2005 with a call for inform ation (70 20 FR 73236) for the development of a revised Integrated Science Assessment for Oxides of 21 Nitrogen and Oxides of Sulfur

- Ecological Criteria (henceforth the "ISA"). An Integrated 22 Review Plan (IRP) was developed to provide the framework and schedule as well as the scope of 23Draft

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18 the review and to identify policy

- relevant questions to be addressed in the components of the 1 review. The IRP was released in 2007 (US EPA , 2007) for CASAC and public review. EPA 2 held a workshop in July 2007 on the I SA to obtain broad input from the relevant scientific 3 communities. This workshop helped to inform the preparation of the first draft ISA , which was 4 released for CASAC and public review in December 2007; a CASAC meeting was held on April 5 2
- 3, 2008 to review the first draft ISA. A second draft ISA was released for CASAC and public 6 review in August 2008, and was discussed at a CASAC meeting held on October 1

- 2, 2008. T he 7 final I SA ( US EPA, 2008) was released in December 2008. 8 Based on the science presented in the I SA , EPA developed a Risk and Exposure 9 Assessment for Review of the Secondary National Ambient Air Quality Standards for Oxides of 10 Nitrogen and Oxides of Sulfur (henceforth the “REA”) to further assess the national impact of 11 the effects documented in the I SA. The Draft Scope and Methods Plan for Risk/Exposure 12 Assessment: Secondary NAAQS Review for Oxides of Nitrogen and Oxides of Sulfur outlining 13 the scope and design of the future REA was prepared for CASAC and public in March 2008. A 14 first draft REA was pre sented to CASAC and the public for review in August 2008 and a second 15 draft was presented for review in June 2009. The final REA (US EPA 2009) was released in 16 September 2009. A first draft Policy Assessment was released in March 2010 and reviewed by 17 CASAC on April 1

- 2, 2010. In a June 22, 2010 letter to the Administrator, CASAC provided 18 advice and recommendations to the Agency concerning the first draft Policy Assessment (Russ ell 19 and Samet, 2010 a). A second draft Policy Assessment was released to CASAC and the public 20 in September 2010 and reviewed b y CASAC on October 6

- 7, 2010. CASAC provided advice 21 and recommendations to the Agency regarding the second draft Policy Assessme nt in a 22 December 9, 2010 letter (Russell and Samet 2010 b ). CASAC and public comments on the 23Draft

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19 second draft Policy

Assessment were considered by EPA staff in developing a final Policy 1 Assessment ( US EPA, 2011). CASAC requested an additional meeting provide additional 2 advice to the Administrator based on the final Policy Assessment on February 15

– 16, 2011. 3 EPA released on January 14, 2011, the final Policy Assessment prior to final document 4 production, to provide sufficient time for CASAC review of the docum ent in advance of this 5 meeting. A final Policy Assessment, incorporating final reference checks and document 6 formatting, was released in February 2011. In a May

17, 2011 letter (Russel l and Samet, 2011 a), 7 CASAC offered additional advice to the Administrat or with regard to recommendations and 8 revisions to the secondary NAAQS for oxides of nitrogen and oxides of sulfur. 9 On February 3, 2006, the Center for Biological Diversity and four other plaintiffs filed 10 an amended complaint alleging that EPA had failed to complete the current review within the 11 period provided b y statute. 2 The schedule for completion of this review i s governed b y a consent 12 decree resolving that lawsuit and the subsequent extension granted to the Agency by the 13 plaintiffs. The schedule presented in the original consent decree that governs this review, entered 14 by the court on November 19, 2007, was revis ed on October 22, 2009 to allow for a 17 month 15 extension of the schedule. The current decree provides that EPA sign for publication notices of 16 proposed and final rulemaking concerning its review of the oxides of nitrogen and oxides of 17 sulfur NAAQS no late r than July 12, 2011 and March 20, 2012, respectively

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

oxides of nitrogen and oxides of sulfur standards. Throughout this preamble a number 20 of conclusion s, findings, and determinations proposed by the Administrator are noted. While 21 they

identify the reasoning that supports this proposal, they are not intended to be final or 22 conclusive in nature. The EPA invites general, specific, and/or technical comment s on all issues 23

2 Center for Biological Diversity, et al. v. Johnson, No. 05

– 1814 (D.D.C.)Draft

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20 involved with this proposal, including all such proposed judgments, conclusions, findings, and 1 determinations. 2 E. Scope of the Current Review 3 In conducting this periodic review of the secondary NAAQS for oxides of nitrogen and 4 oxides of su lfur, as discussed in the IRP and REA, EPA decided to assess the scientific 5 information, associated risks, and standards relevant to protecting the public welfare from 6 adverse effects associated jointly with oxides of nitrogen and sulfur. Although EPA has 7 historically

adopted separate secondary standards for oxides of nitrogen and oxides of sulfur, 8 EPA is conducting a joint review of these standards because oxides of nitrogen and sulfur, and 9 their associated transformation products are linked from an atmos pheric chemistry perspective, 10 as well as from an environmental effects perspective. The National Research Council (NRC) has 11 recommended that EPA consider multiple pollutants, as appropriate, in forming the scientific 12 basis for the NAAQS (NRC, 2004). As d iscussed in the I SA and REA, there is a strong basis for 13 considering these pollutants together, building upon EPA’s past recognition of the interactions of 14 these pollutants and on the growing bod y of scientific information that is now available related to 15 these interactions and associated ecological effects. 16 In defining the scope of this review, it must be considered that EPA has set secondary 17 standards for two other criteria pollutants related to oxides of nitrogen and sulfur: ozone and 18 particulate matte r (PM). Oxides of nitrogen are precursors to the formation of ozone in the 19 atmosphere, and under certain conditions, can combine with atmospheric ammonia to form 20 ammonium nitrate, a component of fine PM. Oxides of sulfur are precursors to the formation o f 21 particulate sulfate, which is a significant component of fine PM in many parts of the U.S. There 22 are a number of welfare effects directl y associated with ozone and fine PM, including ozone

- 23Draft

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21 related damage to vegetation and PM

- related visibility impairment. Protection against those 1 effects is provided b y the ozone and fine PM secondary standards. This review focuses on 2 evaluation of the protection provided b y secondary standards for oxides of nitrogen and sulfur 3 for two general t ypes of effects: (1) dir ect effects on vegetation associated with exposure to 4 gaseous oxides of nitrogen and sulfur in the ambient air, which are the effects that the current 5 NO 2 and SO 2 secondary standards protect against and (2) effects associated with the deposition 6 of oxides of nitrogen and sulfur to sensitive aquatic and terrestrial ecos ys tems, including 7 deposition in the form of particulate nitrate and particulate sulfate. 8 The I SA focuses on the ecological effects associated with deposition of ambient oxides of 9 nitrogen and sulfur to natural sensitive ecos ys tems, as distinguished from commercially managed 10 forests and agricultural lands. This focus reflects the fact that the majority of the scientific 11 evidence regarding acidification and nutrient enrichment is based on studi es in unmanaged 12
ecos ys tems. Non
- managed terrestrial ecos ys tems tend to have a higher fraction of N deposition 13 resulting from atmospheric nitrogen ( US EPA , 2008, section 3.3.2.5). In addition, the I SA notes 14 that agricultural and commercial forest lands ar e routinely fertilized with amounts of nitrogen 15 that exceed air pollutant inputs even in the most polluted areas ( US EPA , 2008, section 3.3.9). 16 This review recognizes that the effects of nitrogen deposition in managed areas are viewed 17 differentl y from a p ublic welfare perspective than are the effects of nitrogen deposition in 18 natural, unmanaged ecosystems, largely due to the more homogeneous, controlled nature of 19 species composition and development in managed ecos ys tems and the potential for benefits of 20 in creased productivity in those ecos ys tems. 21 In focusing on natural sensitive ecosy stems, the Policy Assessment primarily

considers 22 the effects of ambient oxides of nitrogen and sulfur via deposition on multiple ecological 23Draft

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22 receptors. The ISA highlights effe cts including those associated with acidification and nitrogen 1 nutrient enrichment. With a focus on these deposition

- related effects, EPA’s objective is to 2 develop a framework for oxides of nitrogen and sulfur standards that incorporates ecologicall y 3 rele vant factors and that recognizes the interactions between the two pollutants as they deposit to 4 sensitive ecos ys tems. The overarching policy objective is to develop a secondary standard(s) 5 that is based on the ecological criteria described in the I SA and the results of the assessments in 6 the REA, and is consistent with the requirement of the CAA to set secondary standards that are 7 requisite to protect the public welfare from an y known or anticipated adverse effects associated 8 with the presence of these air pollutants in the ambient air. Also consistent with the CAA, this 9 policy

objective necessarily

includes consideration of “variable factors . . . which of themselves 10 or in combination with other factors may alter the effects on public welfare” of the crit eria air 11 pollutants included in this review. 12 In addition, we have chosen to focus on the effects of ambient oxides of nitrogen and 13 sulfur on ecological impacts on sensitive aquatic ecosy stems associated with acidifying 14 deposition of nitrogen and sulfur, which is a transformation product of ambient oxides of 15 nitrogen and sulfur. Based on the information in the I SA, the assessments presented in the REA, 16 and advice from CASAC on earlier drafts of this Policy Assessment (Russell and Samet, 2010a, 17 2010b), and as discussed in detail in the Policy Assessment , we have the greatest confidence in 18 the causal linkages between oxides of nitrogen and sulfur and aquatic acidification effects 19 relative to other deposition

- related effects, including terrestrial acidificatio n and aquatic and 20 terrestrial nutrient enrichment. 21 22Draft

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23 II. Rationale for Proposed Decision on the Adequacy of the Current Secondary 1 Standards 2 D ecisions on retaining or revising the current secondary standards for oxides of nitrogen 3 and sulfur are largely public welfare policy

judgments based on the Administrator’s informed 4 assessment of what constitutes requisite protection against adverse effect s to public welfare. A 5 public welfare policy decision should draw upon scientific information and anal ys es about 6 welfare effects, exposure and risks, as well as judgments about the appropriate response to the 7 range of uncertainties that are inherent in th e scientific evidence and analy ses. The ultimate 8 determination as to what level of damage to ecos ystems and the services provided by those 9
ecos ys tems is adverse to public welfare is not wholly a scientific question, although it is 10 informed b y scientific studies linking e cos ys tem damage to losses in ecosy stem services, and 11 information on the value of those losses of ecos ys tem services. In reaching such decisions, the 12 Administrator seeks to establish standards that are neither more nor less stringent than ne cessary 13 for this purpose. 14 This section presents the rationale for the Administrator’s proposed conclusions with 15 regard to the adequacy of protection and ecological relevance of the current secondary standards 16 for oxides of nitrogen and sulfur. As discussed more full y below, this rationale considered the 17 latest scientific information on ecological effects associated with the presence of oxides of 18 nitrogen and oxides of sulfur in the ambient air. This rationale also takes into account: (1) staff 19 assessments of the most policy

- relevant information in the I SA and staff analy ses of air quality , 20 exposure, and ecological risks, presented more fully in the REA and in the Policy Assessment , 21 upon which staff conclusions on revisions to the secondary oxides of nitrog en and oxides of 22 sulfur standards are based; (2) CASAC advice and recommendations, as reflected in discussions 23Draft

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24 of drafts of the I SA, REA, and Policy Assessment at public meetings, in separate written 1 comments, and in CASAC’s letters to the Administrator; and (3) public comments received 2 during the development of these documents, either in connection with CASAC meetings or 3 separately

. 4 In developing this rationale, EPA has drawn upon an integrative s ynthesis of the entire 5 body of evidence, published through early 2008 , on ecological effects associated with the 6 deposition of oxides of nitrogen and oxides of sulfur in the ambient air (US EPA, 2008)

. As 7 discussed below in section II.A, this body of evidence addresses a broad range of ecological 8 endpoints asso ciated with ambient levels of oxides of nitrogen and oxides of sulfur

. In 9 considering this evidence, EPA focuses on those ecological endpoints, such as aquatic 10 acidification, which the ISA judges associations with oxides of nitrogen and oxides of sulfur to 11 be causal, likel y causal, or for which the evidence is highly suggestive that oxides of nitrogen 12 and/or sulfur contribute to the reported effects. 13 Crucial to this review is the development of a form for an ecologicall y relevant standard 14 that reflects both the geographicall y variable and deposition

- dependent nature of the effects. 15 The atmospheric levels of oxides of nitrogen and sulfur that afford a particular level of 16
ecos ys tem protection are those levels that result in an amount of deposition that is l ess than the 17 amount of deposition that a given ecos ys tem can accept without defined levels of degradation of 18 the ecological indicator for a targeted effect. 19 Drawing from the framework developed in the REA, the framework we used to structure 20 an ecologically meaningful secondary standard in the Policy

Assessment and to further develop 21 the indicator, form, level, and averaging time of such as standard in section III of this proposal is 22Draft

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depicted below and highlights the three key linkages that need to be consid 1 ecologicall y relevant standard. 2 3

4 Figure II - 1. Sim plified conceptual design of the form of an aquatic acidification standard 5 for oxides of nitrogen and sulfur. 6 7 The following discussion 8 The Policy

Assessment includes staff’s evaluation of the policy implications of the scientific 9 assessment of the evidence presented and assessed in the ISA and the results of quantitative 10 assessments based on that information presented 11 information informs staff conclusions and the development of policy options in the 12 Assessment for consideration in addressing public and welfare effects associated with the 13 presence of oxides of nitrogen and oxides of sulfur in the ambient air. Of particular note, chapter 14 2 of the Policy

Assessment presents information not repeated here that characterizes emissions, 15 air quality , deposition and water quality. It includes discussions of the sources of nitr 16 sulfur in the atmosphere as well as current ambient air quality monitoring networks and models. 17 Additional information in this section includes ecological modeling and water qualit y data 18
sources. 19 Section II.A presents a discussion 20 sulfur in the ambient air. The discussion 21 25 depicted below and highlights the three key linkages that need to be consid ered in developing an

Sim plified conceptual design of the form of an aquatic acidification standard for oxides of nitrogen and sulfur.
The following discussion relies heavil y on Chapters 2 and 3 of the Policy Assessment includes staff’s evaluation of the policy implications of the scientific assessment of the evidence presented and assessed in the ISA and the results of quantitative assessments based on that information presented and assessed in the REA. Taken together, this information informs staff conclusions and the development of policy options in the

for consideration in addressing public and welfare effects associated with the

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

discussion of the effects associated with oxides of nitrogen and discussion is organized around the types of effects being June 15, 2011 ered in developing an

Sim plified conceptual design of 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 and assessed in the REA. Taken together, this information informs staff conclusions and the development of policy options in the Policy

for consideration in addressing public and welfare effects associated with the

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 nitr ogen and sulfur in the atmosphere as well as current ambient air quality monitoring networks and models. Additional information in this section includes ecological modeling and water qualit y data
associated with oxides of nitrogen and is organized around the types of effects beingDraft

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26 considered, including direct effects of gaseous oxides of nitrogen and sulfur , deposition

- related 1 effects related to acidification and nutrient enrichment , and other effects such as materials 2 damage, climate

- related effects and mercury methylation

. 3 Section II
. B presents a summary and discussion of the risk and exposure assessment 4 performed for each of the four major ef fects categories. The REA uses case studies representing 5 the broad geographic variability of the impacts from oxides of nitrogen and sulfur to conclude 6 that there are ongoing adverse effects in many ecosy stems from deposition of oxides of nitrogen 7 and sul fur and that under current emissions scenarios these effects are likely to continue. 8 Section II
. C presents a discussion of adversit y related to linking ecological effects to 9 measures that can be used to characterize the extent to which such effects are re asonabl y 10 considered to be adverse to public welfare. This involves consideration of how to characterize 11 adversit y from a public welfare perspective. In so doing, consideration is given to the concept of 12
ecos ys tem services, the evidence of effects on ecos ys tem services, and how ecos ys tem services 13 can be linked to ecological indicators. 14 Section II
. D presents an assessment of the adequacy of the current oxides of nitrogen and 15 oxides of sulfur secondary standards. Consideration is given both to the adequacy of protection 16 afforded b y the current standards for both direct and deposition

- related effects, as well as to the 17 appropriateness of the fundamental structure and the basic elements of the current s tandards for 18 providing protection from deposition

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

- related effects that could reasonably be judged to be adverse to public welfare are 20 occurring under current conditions which are allowed by the current standards is also considered

. 21 Discussion of the structures and basic elements of the current NO 2 and SO 2 secondary standards 22 and the degree to which they are inadequate to protect against such effects is presented

. 23Draft

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27 A. Ecological Effects 1 Thi s section discusses the known or anticipated ecological effects associated with oxides 2 of nitrogen and sulfur, including the direct effects of gas

- phase exposure to oxides of nitrogen 3 and sulfur (section II.A.1) and effects associated with deposition

- relat ed exposure (sections 4 II.A.2 and 3)

. Section II .A. 2 addresses effects related to acidification of aquatic and terrestrial 5
ecos ys tems and s ection II A.3 addresses effects related to nutrient enrichment of aquatic and 6 terrestrial ecos ys tems
. These section s also address questions about the nature and magnitude of 7
ecos ys tem responses to reactive nitrogen and sulfur deposition, including responses related to 8 acidification, nutrient depletion, and, in Section II.A 4 the mobilization of toxic metals in 9 sensitiv e aquatic and terrestrial ecos ys tems . T he uncertainties and limitations associated with the 10 evidence of such effects are also discussed throughout this section

. 11 1. Effects Associated with Gas

-Phase Oxides of Nitrogen and Sulfur 12 E cological effects on vegetation as discussed in earlier reviews as well as the I SA, can 13 be attributed to gas

- phase oxides of nitrogen and sulfur. Acute and chronic exposures to gaseous 14 pollutants such as sulfur dioxide (SO 2 ), nitrogen dioxide (NO 2 ), nitric oxide (NO), nitric acid 15 (HNO 3 ) and perox yacet yl nitrite (PAN) are associated with negative impacts to vegetation. The 16 current secondary

NAAQS were set to protect against direct damage to vegetation by exposure to 17 gas - phase oxides of nitrogen and sulfur , such as foliar in jury , decreased photosy

nthesis, and 18 decreased growth. The following summary is a concise overview of the known or anticipated 19 effects to vegetation caused by gas phase nitrogen and sulfur

. Most phototoxic effects associated 20 with gas phase oxides of nitro gen and sulfur occur at levels well above ambient concentrations 21 observed in the U.S. ( US EPA, 2008, section 3.4.2.4). 22Draft

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28 a. Nature of ecosystem responses to gas

- phase nitrogen and sulfur 1 The 2008 I SA found that gas phase N and S are associated with direct phy t otoxic effects 2 ( US EPA, 2008, section 4.4). The evidence is sufficient to infer a causal relationship between 3 exposure to SO 2 and injury to vegetation ( US EPA, 2008, section 4.4.1 and 3.4.2.1). Acute foliar 4 injury

to vegetation from SO 2 may occur at level s above the current secondary standard (3

- h 5 average of 0.50 ppm). Effects on growth, reduced photosy nthesis and decreased yi eld of 6 vegetation are also associated with increased SO 2 exposure concentration and time of exposure. 7 The evidence is sufficient to infer a causal relationship between exposure to NO, NO 2 8 and PAN and injury

to vegetation ( US EPA, 2008, section 4.4.2 and 3.4.2.2). At sufficient 9 concentrations, NO, NO 2 and PAN can decrease photosy nthesis and induce visible foliar injury 10 to plants. Evi dence is also sufficient to infer a causal relationship between exposure to HNO 3 11 and changes to vegetation ( US EPA, 2008, section 4.4.3 and 3.4.2.3). Phytotoxic effects of this 12 pollutant include damage to the leaf cuticle in vascular plants and disappearance of some 13 sensitive lichen species. 14 b. Magnitude of ecosystem response to gas

- phase nitrogen and sulfur 15 Vegetation in ecos ys tems near sources of gaseous oxides of nitrogen and sulfur or where 16 SO 2 , NO, NO 2 , PAN and HNO 3 are most concentrated are more likel y to be impacted b y these 17 pollutants. Uptake of these pollutants in a plant canopy is a complex process involving 18 adsorption to surfaces (leaves, stems and soil) and absorption into leaves ( US EPA, 2008, section 19 3.4.2). The functional relationship between ambient concentrations of gas phase oxides of 20 nitrogen and sulfur and specific plant response are impacted b y internal factors such as rate of 21 stomatal conductance and plant detoxification mechanisms, and external factors including plant 22Draft

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29 water status, light, temperature, humidity , and pollutant exposure regime ( US EPA, 2008, section 1 3.4.2). 2 Entry

of gases into a leaf is dependent upon phy sical and chemical processes of gas phase 3 as well as to stomatal aperture. The aperture of the stomata is controlled l argely by the 4 prevailing environmental conditions, such as water availability , humidity , temperature, and light 5 intensity . When the stomata are closed, resistance to gas uptake is high and the plant has a very 6 low degree of susceptibility to injury . Mosse s and lichens do not have a protective cuticle barrier 7 to gaseous pollutants or stomata and are generall y more sensitive to gaseous sulfur and nitrogen 8 than vascular plants ( US EPA, 2008, section 3.4.2). 9 The appearance of foliar injury can vary signific antly

across species and growth 10 conditions affecting stomatal conductance in vascular plants ( US EPA, 200 9 , section 6.4.1). For 11 example, damage to lichens from SO 2 exposure include s decrease d photosynthesis and 12 respiration, damage to the algal component of the lichen, leakage of electroly tes, inhibition of 13 nitrogen fixation, decreased K+ absorption, and structural changes. 14 The phy

totoxic effects of gas phase oxides of nitrogen and sulfur are dependent on the 15 exposure concentration and duration and species sensitivity to these pollutants. Effects to 16 vegetation associated with oxides of nitrogen and sulfur are therefore variable across the U.S. 17 and tend to be higher near sources of photochemical smog. For example, SO 2 is considered to be 18 the primary

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

. 19 The I SA states there is very limited new research on phy totoxic effects of NO, NO 2 , PAN 20 and HNO 3 at concentrations currentl y observed in the United States with the exception of som e 21 lichen species ( US EPA, 2008, section 4.4). Past and current HNO 3 concentrations may be 22 contributing to the decline in lichen s pecies in the Los Angeles basin . Most phototoxic effects 23Draft

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30 associated with gas phase oxides of nitrogen and sulfur occur at lev els well above ambient 1 concentrations observed in the U.S. ( US EPA, 2008, section 3.4.2.4). 2 2. Acidification Effects Associated with Deposition of Oxides of Nitrogen and Sulfur 3 Sulfur oxides and nitrogen oxides in the atmosphere undergo a complex mix of reactions 4 in gaseous, liquid, and solid phases to form various acidic compounds. These acidic compounds 5 are removed from the atmosphere through deposition: either wet (e.g., rain, snow), fog or cloud, 6 or dry

(e.g., gases, particles). Deposition of these ac idic compounds to ecosy stems can lead to 7 effects on ecos ys tem structure and function. Following deposition, these compounds can, in 8 some instances unless retained b y soil or biota, leach out of the soils in the form of sulfate (SO 4 2
- 9 ) and nitrate (NO 3 - ), l eading to the acidification of surface waters. The effects on ecosy stems 10 depend on the magnitude and rate of deposition, as well as a host of biogeochemical processes 11 occurring in the soils and waterbodies ( US EPA, 200 9 , section 2.1). The chemical forms of 12 nitrogen that may

contribute to acidifying deposition include both oxidized and reduced chemical 13 species , including NH x . 14 When sulfur or nitrogen leaches from soils to surface waters in the form of SO 4 2
- or NO 3 - , 15 an equivalent amount of positive cations, or countercharge, is also transported. This maintains 16 electroneutrality . I f the countercharge is provided by base cations, such as calcium (Ca2+ ), 17 magnesium (Mg 2+ ), sodium (Na+ ), or potassium (K+ ), rather than h ydrogen (H + ) and dissolved 18 inorganic aluminum , the acidity of the soil water is neutralized, but the base saturation of the soil 19 decreases. Continued SO 4 2
- or NO 3 - leaching can deplete the available base cation pool in soil. 20 As the base cations are removed, continued deposition and leaching of SO 4 2
- and/or NO 3 - (with 21 H + and Al 3+ ) leads to acidification of soil water, and by connection, surface water. Introduction 22 of strong acid anions such as sulfate and nitrate to an already acidic soil, whether naturally or due 23Draft

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31 to anthropogenic activities, can lead to instantaneous acidification of waterbodies through direct 1 runoff without an y significant change in base cation saturation. The ability of a watershed to 2 neutralize acidic deposition is determined by a variety of biogeoph ys ical factors including 3 weathering rates, bedrock composition, vegetation and microbial processes, phy sical and 4 chemical characteristics of soils and hydrologic flowpaths. ( US EPA, 200 9 , section 2.1) Some of 5 these factors such as vegetation and soil depth are highl y variable over small spatial scales such 6 as meters, but can be aggregated to evaluate patterns over larger spatial scales. Acidifying 7 deposition of oxides of nitrogen and sulfur and the chemical and biological responses associated 8 with these inputs vary temporally

. Chronic or long

- term deposition processes in the time scale of 9 ye ars to decades result in increases in inputs of N and S to ecos ys tems and the associated 10 ecological effects. Episodic or short term (i.e., hours or day s) deposition refers to events in which 11 the leve l of the acid neutralizing capacity (ANC) of a lake or stream is temporaril y lowered. In 12 aquatic ecosy stems, short

- term (i.e., hours or day s) episodic changes in water chemistry can have 13 significant biological effects. Episodic acidification refers to co nditions during precipitation or 14 snowmelt events when proportionately more drainage water is routed through upper soil horizons 15 that tend to provide less acid neutralizing than was passing through deeper soil horizons ( US 16 EPA, 200 9 , section 4.2). In addit ion, the accumulated sulfate and nitrate in snow packs can 17 provide a surge of acidic inputs. Some streams and lakes may have chronic or base flow 18 chemistry

that is suitable for aquatic biota, but may be subject to occasional acidic episodes with 19 deleterio us consequences to sensitive biota. 20 The following summary is a concise overview of the known or anticipated effects caused 21 by acidification to ecos ys tems within the United States. Acidification affects both terrestrial and 22 freshwater aquatic ecosystems. 23Draft

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32 a. Nature of acidification

- related ecosystem responses 1 The I SA concluded that deposition of oxides of nitrogen and sulfur and NH x leads to the 2 varying degrees of acidification of ecos ys tems ( US EPA, 2008 ). In the process of acidification, 3 biogeochemical c omponents of terrestrial and freshwater aquatic ecos ys tems are altered in a way 4 that leads to effects on biological organisms. Deposition to terrestrial ecosy stems often moves 5 through the soil and eventually leaches into adjacent water bodies. 6 i. Aquatic eco systems 7 The scientific evidence is sufficient to infer a causal relationship between acidifying 8 deposition and effects on biogeochemistry and biota in aquatic ecos ys tems ( US EPA, 2008, 9 section 4.2.2). The strongest evidence comes from studies of surface w ater chemistry in which 10 acidic deposition is observed to alter sulfate and nitrate concentrations in surface waters, the sum 11 of base cations, ANC, diss olved inorganic aluminum and pH ( US EPA, 2008, section 3.2.3.2). 12 ANC is a key

indicator of acidification with relevance to both terrestrial and aquatic ecosy stems. 13 ANC is useful because it integrates the overall acid

- base status of a lake or stream and reflects 14 how aquatic ecos ys tems respond to acidic deposition over time. There is also a relationship 15 betwee n ANC and the surface water constituents that directl y contribute to or ameliorate acidity

16 related stress, in particular, concentrations of h ydrogen ion (as pH), Ca2+ , and aluminum. 17 Moreover, low pH surface waters leach aluminum from soils, which is quite lethal to fish and 18 other aquatic organisms. In aquatic s ys tems, there is a direct relationship between ANC and fish 19 and phy

to - zooplankton diversity

and abundance

. 20 Low ANC coincides with effects on aquatic s ys tems (e.g., individual species fitness loss 21
o r death, reduced species richness, altered community structure). At the community level, 22 species richness is positively correlated with pH and ANC because energy cost in maintaining 23Draft

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33 phy siological homeostasis, growth, and reproduction is high at low ANC lev els. For example, 1 there is a logistic relationship between fish species richness and ANC class for Adirondack Case 2 Study

Area lakes that indicates the probability of occurrence of an organism for a given value of 3 ANC
. Biota are generally not harmed when A

NC values are >100 microequivalents per liter 4 (μeq/L ). The number of fish species also peaks at ANC values >100 μeq/L. Below 100 μeq/L, 5 ANC fish fitness and community diversity begin to decline ( US EPA, section 4.2) . At ANC 6 levels between 100 and 50 μeq/L , the fitness of sensitive species (e.g., brook trout, zooplankton) 7 also begins to decline. When ANC concentrations are <50 μeq/L , they are generall y associated 8 with death or loss of fitness of biota that are sensitive to acidification. 9 Consistent and coher ent documentation from multiple studies on various species from all 10 major trophic levels of aquatic sy stems shows that geochemical alteration caused by acidification 11 can result in the loss of acid

- sensitive biological species (US EPA, 2008, section 3.2.3.3 ). This is 12 most often discussed with relation to pH. For example, in the Adirondacks, of the 53 fish species 13 recorded in Adirondack lakes about half (26 species) were absent from lakes with pH below 6.0. 14 Biological effects are linked to changes in water chemistry including decreases in ANC and pH 15 and increases in inorganic Al concentration. The direct biological effects are caused b y lowered 16 pH which leads to in increased inorganic Al concentrations ( US EPA, 20 11 , Figures 3

- 1 and 3

- 17 2 ). While ANC level d oes not cause direct biological harm it is a good overall indicator of the 18 risk of acidification ( US EPA, 20 11 , section 3.1.3). 19 There are clear associations between ANC, pH, and aquatic species mortality and health 20 which are summarized in section 3.1.1 of the Policy Assessment
. Significant harm to sensitive 21 aquatic species has been observed at pH levels below 6. Normal stream pH levels with little to 22 no toxicity

ranges from 6 to 7 (MacAvo y et al, 1995). Baker et al (1990) observed that “lakes 23Draft

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34 with pH le ss than approximately 6.0 contain significantly

fewer species than lakes with pH levels 1 above 6.0”. As noted in Chapter 3, ty picall y at pH <4.5 and an ANC <0 μeq/L , complete to 2
near - complete loss of many taxa of organisms occur, including fish and aquatic insect 3 populations, whereas other taxa are reduced to only acidophilic species. 4 Additional evidence can help refine the understanding of effects occurring at pH levels 5 between 4.5 and 6. When pH levels are below 5.6, relativel y lower trout survival rat es were 6 observed in the Shenandoah National Park. In field observations, when pH levels dropped to 5, 7 mortality

rates went to 100 percent. (Bulger et al, 2000). At pH levels ranging from 5.4 to 5.8, 8 cumulative mortality continues to increase. Several studies have shown that trout exposed to 9 water with varying pH levels and fish larvae showed increasing mortality as pH levels decrease. 10 In one study

almost 100 percent mortality was observed at a pH of 4.5 compared to almost 100 11 percent survival at a pH of 6.5. Intermediate pH values (6.0, 5.5) in all cases showed reduced 12 survival compared with the control (6.5), but not by statisticall y significant amounts ( US EPA, 13 2008, section 3.2.3.3). 14 One important indicator of acid stress is increased fish mortali ty . The response of fish to 15 pH is not uniform across species. A number of s yn optic survey s indicated loss of species 16 diversity

and absence of several fish species in the pH range of 5.0 to 5.5. If pH is lower, there is 17 a greater likelihood that more fis h species could be lost without replacement, resulting in 18 decreased richness and diversity . In general, populations of salmonids are not found at pH levels 19 less than 5.0, and smallmouth bass (Micropterus dolomieu) populations are usually not found at 20 pH va lues less than about 5.2 to 5.5. From Table 3

- 1, only one study

showed significant 21 mortality

effects above a pH of 6, while a number of studies showed significant mortality when 22 pH levels are at or below 5.5. 23Draft

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35 The highest pH level for any of the studies reported in the I SA, is 6.0, suggesting that pH 1 above 6.0 is protective against mortalit y effects for most species. Most thresholds are in the 2 range of pH of 5.0 to 6.0, which suggests that a target pH should be no lower than 5.0. 3 Protection against mor tality in some recreationally important species such as lake trout (pH 4 threshold of 5.6) and crappie (pH threshold of 5.5), combined with the evidence of effects on 5 larval and embryo survival suggests that pH levels greater than 5.5 should be targeted to p rovide 6 protection against mortality effects throughout the life stages of fish. 7 Non
- lethal effects have been observed at pH levels as high as 6. A study in the 8 Shenandoah National Park found that the condition factor, a measure of fish health expressed a s 9 fish weight/length3 multiplied by a scaling constant, is positively correlated with stream pH 10 levels, and that the condition factor is reduced in streams with a pH of 6.0 ( US EPA, 2008, 11 section 3.2.3.3). 12 Biodiversity is another indicator of aquatic ecos ys tem health. A key study in the 13 Adirondacks found that lakes with a pH of 6.0 had only half the potential species of fish (27 of 14 53 potential species). There is often a positive relationship between pH and number of fish 15 species, at least for pH values b etween about 5.0 and 6.5, or ANC values between about 0 to 16 100 μeq/L . Such observed relationships are complicated, however, by the tendency

for smaller 17 lakes and streams, having smaller watersheds, to also support fewer fish species, irrespective of 18 acid

- b ase chemistry . This pattern may be due to a decrease in the number of available niches as 19 stream or lake size decreases. Nevertheless, fish species richness is relatively easily determined 20 and is one of the most useful indicators of biological effects of s urface water acidification. 21 Changes in stream water pH and ANC also contribute to declines in taxonomic richness 22 of zooplankton, and macroinvertebrates which are often sources of food for fish, birds and other 23Draft

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36 animal species in various ecos ys tems. These fish may also serve as a source of food and 1 recreation for humans. Acidification of ecos ys tems has been shown to disrupt food web 2 dy namics causing alteration to the diet, breeding distribution and reproduction of certain species 3 of birds ( US EPA, 2008, se ction 4.2.2.2. and Table 3

- 9). For example, breeding distribution of 4 the common goldeney e (Bucephala clangula ) an insectivorous duck may be affected b y changes 5 in acidifying deposition. Similarly , decreases in prey diversit y and quantity have been observ ed 6 to create feeding problems for nesting pairs of loons on low

- pH lakes in the Adirondacks. 7 ii. Terrestrial ecosystems 8 In terrestrial ecos ys tems, the evidence is sufficient to infer a causal relationship between 9 acidifying deposition and changes in biogeochemistry ( US EPA, 2008, section 4.2.1.1). The 10 strongest evidence comes from studies of forested ecos ys tems, with supportive information on 11 other plant taxa, including shrubs and lichens (US EPA, 2008, section 3.2.2.1.). Three useful 12 indicators of chemical changes and acidification effects on terrestrial ecos ystems, showing 13 consistency

and coherence among multiple studies are: soil base saturation, Al concentrations in 14 soil water and soil C:N ratio ( US EPA, 2008, section 3.2.2.2). 15 As discussed in the I SA and REA, i n soils with base saturation less than about 15 to 20%, 16 exchange chemistry is dominated by

Al. Under these conditions, responses to inputs of sulfuric 17 acid and HNO 3 largel y involve the release and mobilization of dissolved inorganic Al. The 18 effect can be neutralized by weathering from geologic parent material or base cation exchange. 19 The Ca2+ and Al concentrations in soil water are strongl y influenced b y soil acidification and 20 both have been shown to have quantitative links to tree health, including Al interference with 21 Ca2+ uptake and Al toxicity to roots. Effects of nitrification and associated acidification and 22Draft

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37 cation leaching have been consistently shown to occur onl y in soils with a C:N ratio below about 1 20 to 25. 2 Soil acidificatio n caused by acidic deposition has been shown to cause decreased growth 3 and increased susceptibility to disease and injury

in sensitive tree species. Red spruce ( Picea 4 rubens ) dieback or decline has been observed across high elevation areas in the Adironda ck, 5 Green and White mountains. The frequency of freezing injury

to red spruce needles has 6 increased over the past 40 y ears, a period that coincided with increased emissions of S and N 7 oxides and increased acidifying deposition. Acidifying deposition can contribute to dieback in 8 sugar maple (

Acer saccharum ) through depletion of cations from soil with low levels of available 9 Ca. Grasslands are likel y less sensitive to acidification than forests due to grassland soils being 10 generall y rich in base cations. 11 iii. Ecosystem sensitivity 12 The intersection between current deposition loading, historic loading, and sensitivity 13 defines the ecological vulnerability to the effects of acidification. Freshwater aquatic and some 14 terrestrial ecos ys tems, notably forests, are the ecosy stem ty pes which are most sensitive to 15 acidification. The IS A reports that the principal factor governing the sensitivity of terrestrial and 16 aquatic ecosy stems to acidification from sulfur and nitrogen deposition is geology (particularly 17 surficial geology). Geologic formations having low base cation supply generall y underlie the 18 watersheds of acid

- sensitive lakes and streams. Other factors that contribute to the sensitivity of 19 soils and surface waters to acidifying deposition include topography , so il chemistry , land use, 20 and hydrologic flowpaths. Episodic and chronic acidification tends to occur in areas that have 21 base

- poor bedrock, high relief, and shallow soils ( US EPA, 2008, section 3.2.4.1). 22Draft

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38 b. Magnitude of acidification

- related ecosystem responses 1 Terrestrial and aquatic ecosy stems differ in their response to acidifying deposition. 2 Therefore the magnitude of ecos ys tem response is described separatel y for aquatic and terrestrial 3
ecos ys tems in the following sections. The magnitude of response refe rs to both the severit y of 4 effects and the spatial extent of the U.S. which is affected. 5 i. Aquatic acidification 6 Freshwater ecos ys tem survey s and monitoring in the eastern United States have been 7 conducted b y man y programs since the mid

- 1980s, including EPA ’s Environmental Monitoring 8 and Assessment Program (EMAP), National Surface Water Survey (NSWS), Temporall y 9 Integrated Monitoring of Ecosy stems (TIME), and L ong - term Monitoring (LTM) programs. 10 Based on anal ys es of surface water data from these programs, Ne w England, the Adirondack 11 Mountains, the Appalachian Mountains (northern Appalachian Plateau and Ridge/Blue Ridge 12 region), and the Upper Midwest contain the most sensitive lakes and streams (i.e., ANC less than 13 about 50 μeq/L ). Portions of northern Florida also contain many acidic and low

- ANC lakes and 14 streams, although the role of acidifying deposition in this region is less clear. The western U.S. 15 contains many

of the surface waters most sensitive to potential acidification effects, but with the 16 exception of the L os Angeles Basin and surrounding areas, the levels of acidifying deposition are 17 low in most areas. Therefore, acidification of surface waters b y acidic deposition is not as 18 prevalent in the western U.S., and the extent of chronic surface water ac idification that has 19 occurred in that region to date has likel y been very limited relative to the Eastern U.S. ( US EPA, 20 2008, section 3.2.4.2 and US EPA, 200 9 , section 4.2.2). 21 There are a number of species including fish, aquatic insects, other invertebra tes and 22 algae that are sensitive to acidification and cannot survive, compete, or reproduce in acidic 23Draft

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39 waters ( US EPA, 2008, section 3.2.3.3). Decreases in ANC and pH have been shown to 1 contribute to declines in species richness and declines in abundance of zooplankton, 2 macroinvertebrates, and fish. Reduced growth rates have been attributed to acid stress in a 3 number of fish species including Atlantic salmon (Salmo salar), Chinook salmon (Oncorh ynchus 4 tshawy

tscha), lake trout (Salvelinus namaycush), rainbow trout (Oncorh ynchis my kiss), brook 5 trout (Salvelinus Fontinalis), and brown trout (Salmo trutta). In response to small to moderate 6 changes in acidity , acid

- sensitive species are often replaced by other more acid

- tolerant species, 7 resulting in changes in c ommunity composition and richness. The effects of acidification are 8 continuous, with more species being affected at higher degrees of acidification. At a point, 9 ty picall y a pH <4.5 and an ANC <0 μeq/L , complete to near

- complete loss of many taxa of 10 organ isms occur, including fish and aquatic insect populations, whereas other taxa are reduced to 11 only

acidophilic species. These changes in taxa composition are associated with the high energy 12 cost in maintaining ph ys iological homeostasis, growth, and reproduc tion at low ANC levels ( US 13 EPA, 2008, section 3.2.3.3) . Decreases in species richness related to acidification have been 14 observed in the Adirondack Mountains and Catskill Mountains of New York, New England and 15 Pennsylvania, and Virginia. From the sensitive areas identified by the ISA, further “case study” 16 anal ys es on aquatic ecosy stems in the Adirondack Mountains and Shenandoah National Park 17 were conducted to better characterize ecological risk associated with acidification (US EPA, 18 200 9 , section 4). 19 ANC is the most widely used indicator of acid sensitivity and has been found in various 20 studies to be the best single indicator of the biological response and health of aquatic 21 communities in acid

- sensitive sy stems (Lien et al., 1992; Sullivan et al., 2006; US EPA, 2008). 22 In the REA, surface water trends in SO 4 2
- and NO 3 - concentrations and ANC levels were 23Draft

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40 anal yz ed to affirm the understanding that reductions in deposition could influence the risk of 1 acidification. ANC values have been categorized according to their effects on biota, as shown in 2 the table below . Monitoring data from TIME/LTM and EMAP programs were assessed for the 3 ye ars 1990 to 2006, and past, present, and future water qualit y levels were estimated by both 4 steady

- state and d yn amic biogeochemical models. 5Draft

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41 Table II - 1. Ecological effects associated with alternative levels of acid neutralizing capacity 1 (ANC). (source: USEPA, Acid Rain Program) 2 Category

L abel ANC Levels and Expected Ecological Effects
Acute Concern <0 μeq/L Complete loss of fish populations is expected. Planktonic communities have extremel y low diversity and are dominated by

acidophilic taxa. The numbers of individuals in plankton species that are present are greatl y reduced. Severe Concern 0

–20 μeq/L High ly sensitive to episodic acidification. During episodes of high acidifying deposition, brook trout populations may experience lethal effects. The diversit y and distribution of zooplankton communities decline sharpl y. Elevated Concern 20

– 50 μeq/L Fish species richness is greatl y 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 zoop lankton 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 Studies on fish species richness in the Adirondacks Case Study Area demonstrated the 4 effect of acidification. Of the 53 fish species recorded in Adirondack Case Study Area lakes, 5 only 27 species were found in lakes with a pH <6.0. The 26 species missing from lakes with a 6 pH <6.0 include important recreational species, such as Atlantic salmon, tiger trout (Salmo trutta 7 X Salvelinus fontinalis ), red breast sunfish (

Lepomis auritus ), bluegill ( Lepomis macrochirus ), 8Draft

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42 tiger musky (

Esox masquinongy X lucius ), walleye ( Sander vitreus ), alewife ( Alosa 1 pseudoharengus ), and kokanee ( Oncorhynchus nerka ), as well as ecologically important 2 minnows that are common ly consumed b y sport fish. A survey of 1,469 lakes in the late 1980s 3 found 346 lakes to be devoid of fish. Among lakes with fish, there was a relationship between 4 the number of fish species and lake pH, ranging from about one species per lake for lakes hav ing 5 a pH <4.5 to about six species per lake for lakes having a pH >6.5. In the Adirondacks, a 6 positive relationship exists between the pH and ANC in lakes and the number of fish species 7 present in those lakes ( US EPA, 2008, section 3.2.3.4). 8 Since the mi d

- 1990s, streams in the Shenandoah Case Study Area have shown slight 9 declines in NO 3 - and SO 4 2
- concentrations in surface waters. The 2006 concentrations are still 10 above pre

- acidification (1860) conditions. MAGIC modeling predicts surface water 11 concentrations of NO 3 - and SO 4 2
- are10
- and 32

- fold higher, respectively , in 2006 than in 1860. 12 The estimated average ANC across 60 streams in the Shenandoah Case Study Area is 57.9 μeq/L 13 (± 4.5 μeq/L ). 55% of all monitored streams in the Shenandoah Case Study Area have a current 14 risk of Elevated, Severe, or Acute. Of the 55%, 18% are chronically acidic today ( US EPA, 15 200 9 , section 4.2.4.3). 16 Based on a deposition scenario for this stud y area that maintains current emission levels 17 from 2020 to 2050, the simulation forecast indicates that a large number of streams still have 18 Elevated to Acute problems with acid ity . In fact, from 2006 to 2050, the percentage of streams 19 with Acute Concern is predicted to increase b y 5%, while the percentage of streams in Moderate 20 Concern decreases b y 5%. 21 Biological effects of increased acidification documented in the Shenandoah C ase Study 22 Area include a decrease in the condition factor in blacknose dace and a decrease in fish 23Draft

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43 biodiversity

associated with decreasing stream ANC. On average, the fish species richness is 1 lower b y one fish species for every 21 μeq/L decrease in ANC in Shenandoah National Park 2 streams ( US EPA, 2008, section 3.2.3.4). 3 ii. Terrestrial acidification 4 The I SA identified a variety of indicators that can be used to measure the effects of 5 acidification in soils. Most effects of terrestrial acidification are obse rved in sensitive forest 6
ecos ys tem in the U.S. Tree health has been linked to the availability of base cations (Bc) in soil 7 (such as Ca2+ , Mg 2+ and potassium), as well as soil Al content. Tree species show a range of 8 sensitivities to Ca/Al and Bc/Al soil m olar ratios, therefore these are good chemical indicators 9 because they

directly relate to the biological effects. Critical Bc/Al molar ratios for a large 10 variet y of tree species ranged from 0.2 to 0.8. This range is similar to critical ratios of Ca/Al. 11 Pl ant toxicity

or nutrient antagonism was reported to occur at Ca/Al molar ratios ranging from 12 0.2 to 2.5 ( US EPA, 200 9 ). 13 There has been no s ys tematic national survey of terrestrial ecos ys tems to determine the 14 extent and distribution of terrestrial ecos ys tem sensitivity to the effects of acidifying deposition. 15 However, one preliminary national evaluation estimated that ~15% of forest ecosy stems in the 16 U.S. exceed the estimated critical load based on soil ANC leaching for S and N deposition by 17 >250 eq ha

- 1 yr - 1 (McNulty et al., 2007). Forests of the Adirondack Mountains of New York, 18 Green Mountains of Vermont, White Mountains of New Hampshire, the Alleghen y Plateau of 19 Pennsylvania, and high

- elevation forest ecos ys tems in the southern Appalachians are the reg ions 20 most sensitive to terrestrial acidification effects from acidifying deposition ( US EPA, 2008, 21 section 3.2.4.2). While studies show some recovery of surface waters, there are widespread 22 measurements of ongoing depletion of exchangeable base cations in forest soils in the 23Draft

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44 northeastern U.S. despite recent decreases in acidifying deposition, indicating a slow recovery 1 time. 2 In the REA, a critical load anal ys is was performed for sugar maple and red spruce forests 3 in the eastern United States by using Bc/Al ratio in acidified forest soils as an indicator to assess 4 the impact of nitrogen and sulfur deposition on tree health. These are the two most commonly 5 studied tree species in North America for effects of acidification. At a Bc/Al ratio of 1.2, red 6 spruce growth can be decreased b y 20%. Sugar maple growth can be decreased by 20% at a 7 Bc/Al ratio of 0.6 ( U S EPA, 2009 , section 4.4). The REA anal ys is determined the health of at 8 least a portion of the sugar maple and red spruce growing in the United States may have been 9 compromised with acidifyingtotal nitrogen and sulfur deposition. Specifically , total nitrogen and 10 sulfur deposition levels exceeded three selected critical loads for tree growth in 3% to 75% of all 11 sugar maple plots across 24 states. For red spr uce, total nitrogen and sulfur deposition levels 12 exceeded three selected critical loads in 3% to 36% of all red spruce plots across eight states ( US 13 EPA, 200 9 , section 4.4). 14
c. Key uncertainties associated with acidification 15 There are different levels of uncertainty associated with relationships between deposition, 16 ecological effects and ecological indicators. In Chapter 7 of the REA, the case study analy ses 17 associated with each targeted effect area were s ynthesized by identifying the strengths, 18 limitation s, and uncertainties associated with the available data, modeling approach, and 19 relationship between the selected ecological indicator and atmospheric deposition as described 20 by the ecological effect function ( US EPA, 200 9 , Figure 1

- 1). A further discussi on of 21 uncertaint y in aquatic and terrestrial ecos ys tems is presented below. The key uncertainties were 22Draft

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45 characterized as follows to evaluate the strength of the scientific basis for setting a national 1 standard to protect against a given effect (US EPA, 200 9 , section 7): 2 (1) Data Availabilit y: high, medium or low quality . This criterion is based on the availability and 3 robustness of data sets, monitoring networks, availability of data that allows for extrapolation 4 to larger assessment areas, and input parameters for modeling and developing the ecological 5 effect function. The scientific basis for the ecological indicator selected is also incorporated 6 into this criterion. 7 (2) Modeling Approach: high, fairl y high, intermediate, or low confidence. This value is based 8 on the strengths and limitations of the models used in the anal ys is and how accepted they are 9 by

the scientific community for their application in this analy sis. 10 (3) Ecological Effect Function: high, fairl y high, intermediate, or low confidence. This ranking is 11 ba sed on how well the ecological effect function describes the relationship between 12 atmospheric deposition and the ecological indicator of an effect. 13 i. Aquatic acidification 14 The REA concludes that the available data are robust and considered high quality . Th ere 15 is high confidence about the use of these data and their value for extrapolating to a larger 16 regional population of lakes. The EPA TIME/LTM network represents a source of long

- term, 17 representative sampling. Data on sulfate concentrations, nitrate con centrations and ANC from 18 1990 to 2006 used for this analy sis as well as EPA EMAP and REMAP survey s, provide 19 considerable data on surface water trends. 20 There is fairly high confidence associated with modeling and input parameters. 21 Uncertainties in water q uality estimates (i.e., ANC) from MAGIC were derived from multiple 22 site calibrations. The 95% confidence interval for pre

- acidification of lakes was an average of 15 23Draft

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46 μeq/Ldifference in ANC concentrations or 10% and 8 μeq/Lor 5% for streams ( US EPA, 200 9 , 1 section 7.1.2). The use of the critical load model used to estimate aquatic critical loads is limited 2 by the uncertainties associated with runoff and surface water measurements and in estimating the 3 catchment suppl y of base cations from the w eathering of bedrock and soils (McNulty et al., 4 2007). To propagate uncertainty in the model parameters, Monte Carlo methods were employ ed 5 to develop an inverse function of exceedances. There is high confidence associated with the 6 ecological effect function develope d for aquatic acidification. In calculating the ANC function, 7 the depositional load for N or S is fixed by the deposition of the other, so deposition for either 8 will never be zero ( US EPA, 200 9 , Figure 7.1

- 6). 9 ii. Terrestrial acidification 10 The available data used to quantify the targeted effect of terrestrial acidification are 11 robust and considered high qualit y. The USFS

- Kane Experimental Forest and significant 12 amounts of research work in the Alleghen y Plateau have produced extensive, peer

- reviewed data 13 sets. A meta
- analy sis of laboratory studies showed that tree growth was decreased by 20% 14 relative to controls for BC/Al ratios ( US EPA, 2008, section 7.2.1 and Figure 7.2

- 1). Sugar 15 maple and red spruce were the focus of the REA since they are demonstrat ed to be negativel y 16 affected b y soil available Ca2+ depletion and high concentrations of available Al, and occur in 17 areas that receive high acidifying deposition, There is high confidence about the use of the REA 18 terrestrial acidification data and their va lue for extrapolating to a larger regional population of 19 forests. 20 There is high confidence associated with the models, input parameters, and assessment of 21 uncertaint y used in the case study for terrestrial acidification. The Simple Mass Balance (SMB) 22
mo del, a commonly used and widel y applied approach for estimating critical loads, was used in 23Draft

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47 the REA anal ys is ( US EPA, 2008, section 7.2.2). There is fairly high confidence associated with 1 the ecological effect function developed for terrestrial acidificat ion ( US EPA, 200 9 , 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 ecosy stems within th e United States. Nutrient

- enrichment 6 affects terrestrial, freshwater and estuarine ecos ys tems. Nitrogen deposition is a major source of 7 anthropogenic nitrogen. For man y terrestrial and freshwater ecos ys tems 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 (NH x ) contribute to nitrogen deposition. For the most part, nitrogen effects on 12
ecos ys tems 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 I SA found that deposition of nitrogen, including oxides of nitrogen and NH x , leads to 16 the nitrogen enrichment of ecos ys tems ( US EPA 2008). In the process of nitrogen enrichment, 17 biogeochemic al components of terrestrial and freshwater aquatic ecos ys tems are altered in a way 18 that leads to effects on biological organisms. 19 i. Aquatic ecosystems 20 In freshwater ecos ys tems, the evidence is sufficient to infer a causal relationship between 21 N depositio n and the alteration of biogeochemical cycling in freshwater aquatic ecos ys tems ( US 22 EPA, 2008, section 3.3.2.3 ). N deposition is the main source of N enrichment to headwater 23Draft

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48 streams, lower order streams and high elevation lakes. The most common chemical indicators 1 that were studied included NO 3 − and dissolved inorganic nitrogen (DIN) concentration in surface 2 waters as well as Chl a:total P ratio. Elevated surface water NO 3 − concentrations occur in both 3 the eastern and western U.S. Studies report a signifi cant correlation between N deposition and 4 lake biogeochemistry by identifying a correlation between wet deposition and [DI N] 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 i n 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
ecos ys tems ( US EPA, 2008, section 3.3.5.3 ). Increas ed N deposition can cause a shift in 10 community

composition and reduce algal biodiversity , especially in sensitive oligotrophic lakes. 11 In the IS A, the evidence is sufficient to infer a causal relationship between Nr deposition 12 and the biogeochemical cyclin g 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 currentl y receive high 14 levels of nitrogen input from human activities ( US EPA, 200 9 , section 5.1.1 ). It is unknown i f 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, 200 9 , section 5.1.1 ). 18 The evidence is suffi cient to infer a causal relationship between N deposition and the 19 alteration of species richness, species composition and biodiversity in estuarine ecos ys tems ( US 20 EPA, 2008, section 4.3.4.2 and 3.3.5.4 ). Atmospheric and non

- atmospheric sources of N 21 contri bute to increased phy toplankton and algal productivity , leading to eutrophication. Shifts in 22Draft

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49 community

composition, reduced h ypolimnetic DO, decreases in biodiversity , and mortalit y of 1 submerged aquatic vegetation are associated with increased N deposition in estuarine s ys tems. 2 ii. Terrestrial Ecosystems 3 The evidence is sufficient to infer a causal relationship between N deposition and the 4 alteration of biogeochemical cycling in terrestrial ecos ys tems ( US EPA, 2008, section 4.3.1.1 5 and 3.3.2.1 ). This is supported by numerous observational, deposition gradient and field addition 6 experiments in sensitive ecos ys tems. T he leaching of NO 3 - in soil drainage waters and the export 7 of NO 3 - in str eam water w ere identified as two of the primary indictors of N enrichment. Several 8 N
- addition studies indicate that NO 3 - leaching is induced b y chronic additions of N

. Studies 9 identified in the I SA found that surface water NO 3 - concentrations exceeded 1 μeq/Lin 10 watersheds receiving about 9 to 13 kg N/ha/ yr of atmospheric N deposition. N deposition 11 disrupts the nutrient balance of ecos ys tems with numerous biogeochemical effects. The chemical 12 indicators that are t ypically measured include NO 3 − leaching, soil C:N ratio, rates of N 13 mineralization, nitrification, de nitrification, foliar N concentration, and soil water NO 3 − and 14 NH 4 + concentrations. Note that N saturation (N leaching from ecos ys tems) does not need to 15 occur to cause effects. Substantial leaching of NO 3 − from forest soils to stream water can acidify 16 downstream waters, leading to effects described in the previous section on aquatic acidification. 17 Due to the complexity of interactions between the N and C cy cling, the effects of N on C budgets 18 (quantified input and output of C to the ecosy stem) are variable . Regional trends in net 19
ecos ys tem productivity (NEP) of forests (not managed for silviculture) have been estimated 20 through models based on gradient studies and meta

- anal ys is. Atmospheric N deposition has been 21 shown to cause increased litter accumulation and carbon storage in above

- ground wood y 22Draft

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50 biomass. In the West, this has lead to increased susceptibility to more severe fires. Less is 1 known regarding the effects of N deposition on C budgets of non

- forest ecosy stems. 2 The evidence is sufficient to infer a causal relationship between N deposition on the 3 alteration of species richness, species composition and biodiversity in terrestrial ecos ys tems ( US 4 EPA, 2008, section 4.3.1.2 ). Some organisms and ecos ys tems are more sensitive to N deposition 5 and effects of N deposition are not observed in all habitats. The most sensitive terrestrial taxa to 6 N deposition are lichens. Empirical evidence indicates that lichens in the U.S. are affected by 7 deposition levels as low as 3 kg N/ha/ yr. Alpine ecosystems are also sen sitive to N deposition, 8 changes in an individual species ( Carex rupestris ) were estimated to occur at deposition levels 9 near 4 kg N /ha/y

r and modeling indicates that deposition levels near 10 kg N/ha/ yralter plant 10 community

assemblages. In several grassl and ecosy stems, reduced species diversity and an 11 increase in non

- native, invasive species are associated with N deposition. 12 iii. Ecosystem sensitivity to nutrient enrichment 13 The numerous ecos ys tem ty pes that occur across the U.S. have a broad range of 14 sensiti vity

to N deposition ( U S EPA, 2008, Table 4

- 4 ). Increased deposition to N

- limited 15
ecos ys tems can lead to production increases that may be either beneficial or adverse depending 16 on the sy stem and management goals. 17 Organisms in their natural environment are commonly adapted to a specific regime of 18 nutrient availability . Change in the availability of one important nutrient, such as N, may result 19 in an imbalance in ecological stoichiometry , with effects on ecos ys tem processes, structure and 20 function. In ge neral, N deposition to terrestrial ecosy stems causes accelerated growth rates in 21 some species deemed desirable in commercial forests but may lead to altered competitive 22 interactions among species and nutrient imbalances, ultimately affecting biodiversity

. The onset 23Draft

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51 of these effects occurs with N deposition levels as low as 3 kg N/ha/ yr in sensitive terrestrial 1
ecos ys tems to N deposition. In aquatic ecos ys tems, N that is both leached from the soil and 2 directly

deposited to the water surface can pollute the s urface water. This causes alteration of the 3 diatom community

at levels as low as 1.5 kg N/ha/y r in sensitive freshwater ecos ys tems. 4 The degree of ecos ys tem effects lies at the intersection of N loading and N

- sensitivity

. 5 N
- sensitivity

is predominately

driven b y the degree to which growth is limited by nitrogen 6 availability . Grasslands in the western United States are t ypically N

- limited ecos ys tems 7 dominated by

a diverse mix of perennial forbs and grass species. A meta

- analy sis discussed in 8 the I SA ( US EP A, 2008, section 3.3.3), indicated that N fertilization increased aboveground 9 growth in all non

- forest ecosy stems except for deserts. In other words, almost all terrestrial 10
ecos ys tems are N
- limited and will be altered b y the addition of anthropogenic nitrogen. 11 Likewise, a freshwater lake or stream must be N

- limited to be sensitive to N

- mediated 12 eutrophication. There are many examples of fresh waters that are N

- limited or N and 13 phosphorous (P) co

- limited ( US EPA, 2008, section 3.3.3.2 ). A large dataset meta - analy sis 14 discussed in the I SA ( US EPA, 2008, section 3.3.3.2), found that N

- limitation occurred as 15 frequentl y as P
- limitation in freshwater ecos ys tems. Additional factors that govern the 16 sensitivity

of ecos ys tems to nutrient enrichment from N deposi tion include rates and form of N 17 deposition, elevation, climate, species composition, plant growth rate, length of growing season, 18 and soil N retention capacity ( US EPA, 2008, section 4.3 ). Less is known about the extent and 19 distribution of the terrestrial ecos ys tems in the U.S. that are most sensitive to the effects of 20 nutrient enrichment from atmospheric N deposition compared to acidification. 21 Because the productivity of estuarine and near shore marine ecos ys tems is generall y 22 limited by

the availability of N, they are susceptible to the eutrophication effect of N deposition 23Draft

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52 ( US EPA, 2008, section 4.3.4.1 ). A recent national assessment of eutrophic conditions in 1 estuaries found the most eutrophic estuaries were generall y those that had large watershed

- to 2
e stuarine surface area, high human population density , high rainfall and runoff, low dilution, and 3 low flushing rates. In the REA, the National Oceanic and Atmospheric Administration’s 4 (NOAA) National Estuarine Eutrophication Assessment (NEEA) assessment t ool, Assessment of 5 Estuarine Tropic Status (ASSETS) categorical Eutrophication Index (EI) was used to evaluate 6 eutrophication due to atmospheric loading of nitrogen. ASSETS EI is an estimation of the 7 likelihood that an estuary is experiencing eutrophicati on or will experience eutrophication based 8 on five ecological indicators: chlorophyll a , macroalgae, dissolved oxygen, nuisance/toxic algal 9 blooms and submerged aquatic vegetation (SAV). 10 In the REA, two regions were selected for case study anal ys is using ASSETS EI , the 11 Chesapeake Bay

and Pamlico Sound. Both regions received an ASSETS EIrating of Bad 12 indicating that the estuary had moderate to high pressure due to overall human influence and a 13 moderate high to high eutrophic condition ( US EPA, 200 9 , sect ion s 5.2.4.1 and 5.2.4.2 ). These 14 results were then considered with SPAtially Referenced Regression (SPARROW) modeling to 15 develop a response curve to examine the role of atmospheric nitrogen deposition in achieving a 16 desired decrease in load. To change th e Neuse River Estuary ’s EI score from Bad to Poor not 17 only

must 100% of the total atmospheric nitrogen deposition be eliminated, but considerabl y 18 more nitrogen from other sources as well must be controlled (US EPA, 200 9 , section 5.2.7.2). In 19 the Potomac R iver estuary, a 78% decrease of total nitrogen could move the EI score from Bad 20 to Poor ( US EPA, 200 9 , section 5.2.7.1 ). The results of this analy sis indicated decreases in 21 atmospheric deposition alone could not eliminate coastal eutrophication problems due to 22 multiple non

- atmospheric nitrogen inputs ( US EPA, 200 9 , section 7.3.3 ). However, by 23Draft

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53 decreasing atmospheric contributions, it may help avoid the need for more costly controls on 1 nitrogen from other sources. In addition, the somewhat arbitrary discret eness of the EI scale can 2 mask the benefits of decreases in nitrogen between categories. 3 In general, estuaries tend to be N

- limited, and many currentl y receive high levels of N 4 input from human activities to cause eutrophication. As reported in the IS A ( U S EPA, 2008, 5 section 3.2.2.2), atmospheric N loads to estuaries in the U.S. are estimated to range from 2

- 8% 6 for Guadalupe Bay , TX on the lowest end to as high as 72% for St Catherines

- Sapelo estuary , 7 GA. The Chesapeake Bay is an example of a large, well

- s tudied and severely eutrophic estuary 8 that is calculated to receive as much as 30% of its total N load from the atmosphere. 9 b. Magnitude of ecosystem responses 10 i. Aquatic ecosystems 11 The magnitude of ecos ys tem response may be thought of on two time scales, curr ent 12 conditions and how ecosystems have been altered since the onset of anthropogenic N deposition. 13 As noted previousl y, studies found that N

- limitation occurs as frequently as P
- limitation in 14 freshwater ecos ys tems ( US EPA, 2008, section 3.3.3.2 ). Recentl y , a comprehensive stud y of 15 available data from the northern hemisphere surveys of lakes along gradients of N deposition 16 show increased inorganic N concentration and productivity to be correlated with atmospheric N 17 deposition. The results are unequivocal ev idence of N limitation in lakes with low ambient inputs 18 of N, and increased N concentrations in lakes receiving N solel y from atmospheric N deposition. 19 It has been suggested that most lakes in the northern hemisphere may have originall y been N

20 limited, and that atmospheric N deposition has changed the balance of N and P in lakes. 21 Available data suggest that the increases in total N deposition do not have to be larg e to 22 elicit an ecological effect. For example, a hindcasting exercise determined that the change in 23Draft

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54 Rocky Mountain National Park lake algae that occurred between 1850 and 1964 was associated 1 with an increase in wet N deposition that was only about 1.5 kg N/ha. Similar changes inferred 2 from lake sediment cores of the Beartooth Mountains of Wyoming also occurred at about 1.5 kg 3 N/ha deposition. Pre

- industrial inorganic N deposition is estimated to have been onl y 0.1 to 0.7 4 kg N/ha based on measurements from r emote parts of the world. In the western U.S., pre

5 industrial, or background, inorganic N deposition was estimated b y to range from 0.4 to 0.7 kg 6 N/ha/y

r. 7 Eutrophication effects from N deposition are most likely to be manifested in undisturbed, 8 low nutrie nt surface waters such as those found in the higher elevation areas of the western U.S. 9 The most severe eutrophication from N deposition effects is expected downwind of major urban 10 and agricultural centers. High concentrations of lake or streamwater NO 3 − , indicative of 11
ecos ys tem saturation, have been found at a variety of locations throughout the U.S., including the 12 San Bernardino and San Gabriel Mountains within the L os Angeles Air Basin, the Front Range 13 of Colorado, the Allegheny mountains of West Virgini a, the Catskill Mountains of New York, 14 the Adirondack Mountains of New York, and the Great Smoky Mountains in Tennessee ( US 15 EPA, 2008, section 3.3.8 ). 16 In contrast to terrestrial and freshwater s ys tems, atmospheric N load to estuaries 17 contributes to the to tal load but does not necessarily drive the effects since other combined 18 sources of N often greatly exceed N deposition. In estuaries, N

- loading from multiple 19 anthropogenic and non

- anthropogenic pathway s leads to water quality deterioration, resulting in 20 numerous effects including h ypoxic zones, species mortality , changes in community composition 21 and harmful algal blooms that are indicative of eutrophication. The following summary is a 22Draft

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55 concise overview of the known or anticipated effects of nitrogen enric hment on estuaries within 1 the United States. 2 There is a scientific consensus ( US EPA, 2008, section 4.3.4) that nitrogen

- driven 3 eutrophication in shallow estuaries has increased over the past several decades and that the 4 environmental degradation of coast al ecos ys tems due to nitrogen, phosphorus, and other inputs is 5 now a widespread occurrence. For example, the frequency of ph yt oplankton blooms and the 6 extent and severity

of hypoxia have increased in the Chesapeake Bay and Pamlico estuaries in 7 North Carol ina and along the continental shelf adjacent to the Mississippi and Atchafalay a 8 rivers’ discharges to the Gulf of Mexico. 9 A recent national assessment of eutrophic conditions in estuaries found that 65% of the 10 assessed s ys tems had moderate to high overal l eutrophic conditions and generall y received the 11 greatest N loads from all sources, including atmospheric and land

- based sources. Most eutrophic 12 estuaries occurred in the mid

- Atlantic region and the estuaries with the lowest degree of 13 eutrophication were in the North Atlantic. Other regions had mixtures of low, moderate, and high 14 degrees of eutrophication ( US EPA, 2008, section 4.3.4.3 ). 15 The mid

- Atlantic region is the most heavily impacted area in terms of moderate or high 16 loss of submerged aquatic veget ation due to eutrophication ( US EPA, 2008, section 4.3.4.2 ). 17 Submerged aquatic vegetation is important to the quality of estuarine ecosy stem habitats because 18 it provides habitat for a variet y of aquatic organisms, absorbs excess nutrients, and traps 19 sedim ents ( US EPA, 2008, section 4.3.4.2). It is partly because man y estuaries and near

- coastal 20 marine waters are degraded by nutrient enrichment that they are highl y sensitive to potential 21 negative impacts from nitrogen addition from atmospheric deposition. 22Draft

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56 ii. T errestrial ecosystems 1 Little is known about the full extent and distribution of the terrestrial ecos ystems in the 2 U.S. that are most sensitive to impacts caused b y nutrient enrichment from atmospheric N 3 deposition. As previously stated, most terrestrial e cos ys tems are N
- limited, therefore they are 4 sensitive to perturbation caused b y N additions ( US EPA, 2008, section 4.3.1) . Effects are most 5 likely

to occur where areas of relativel y high atmospheric N deposition intersect with N

- limited 6 plant communities. The alpine ecos ys tems of the Colorado Front Range, chaparral watersheds of 7 the Sierra Nevada, lichen and vascular plant communities in the San Bernardino Mountains and 8 the Pacific Northwest, and the southern California coastal sage scrub (CSS) community are 9 among the most sensitive terrestrial ecos ys tems. There is growing evidence ( US EPA, 2008, 10 section 4.3.1.2) that existing grassland ecos ys tems in the western United States are being altered 11 by elevated levels of N inputs, including inputs from atmospheri c deposition. 12 In the eastern U.S., the degree of N saturation of the terrestrial ecos ys tem is often 13 assessed in terms of the degree of NO 3 − leaching from watershed soils into ground water or 14 surface water. Studies have estimated the number of surface wate rs at different stages of 15 saturation across several regions in the eastern U.S. Of the 85 northeastern watersheds examined 16 60% were in Stage 1 or Stage 2 of N saturation on a scale of 0 (background or pretreatment) to 3 17 (visible decline). Of the northeaste rn sites for which adequate data were available for assessment, 18 those in Stage 1 or 2 were most prevalent in the Adirondack and Catskill Mountains. Effects on 19 individual plant species have not been well studied in the U.S. More is known about the 20 sensitivi ty

of particular plant communities. Based largel y on results obtained in more extensive 21 studies conducted in Europe, it is expected that the more sensitive terrestrial ecos ys tems include 22Draft

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57 hardwood forests, alpine meadows, arid and semi

- arid lands, and grass land ecos ys tems ( US 1 EPA, 2008, section 3. 3.5)

. 2 The REA used published research results ( US EPA, 200 9 , section 5.3.1 and US EPA, 3 2008, Table 4.4) to identify meaningful ecological benchmarks associated with different levels of 4 atmospheric nitrogen deposition. These are illustrated in Figure 3

- 4 of the Policy Assessment
. 5 The sensitive areas and ecological indicators identified by the IS A were analyzed further in the 6 REA to create a national map that illustrates effects observed from ambient and experimental 7 atmospheric nitrogen deposition loads in relation to Community Multi

- scale Air Qualit y 8 (CMAQ) 2002 modeling results and NADP monitoring data. T his map, reproduced in Figure 3

- 9 5 of the Policy

Assessment , depicts the sites where empirical effects of terrestrial nutrient 10 enrichment have been observed and site proximity to elevated atmospheric N deposition. 11 Based on information in the I SA and initi al anal ys is in the REA, further case study 12 anal ys es on terrestrial nutrient enrichment of ecosy stems were developed for the CCS 13 community

and Mixed Conifer Forest (MCF) ( US EPA, 2009). Geographic information sy stems 14 (GI S) analy sis supported a qualitative review of past field research to identify ecological 15 benchmarks associated with CSS and mycorrhizal communities, as well as MCF’s nutrient

16 sensitive acidoph yt e lichen communities, fine

- root biomass in Ponderosa pine, and leached 17 nitrate in receiving waters

. 18 The ecological benchmarks that were identified for the CSS and the MCF are included in 19 the suite of benchmarks identified in the I SA ( US EPA, 2008, section 3.3 ). There are sufficient 20 data to confidentl y relate the ecological effect to a loading of atmo spheric nitrogen. For the CSS 21 community , the following ecological benchmarks were identified: 22Draft

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58 (1) 3.3 kg N/ha/ yr
– the amount of nitrogen uptake b y a vigorous stand of CSS; above this level, 1 nitrogen may

no longer be limiting 2 (2) 10 kg N/ha/ yr
– mycorrhizal commun ity changes 3 For the MCF community, the following ecological benchmarks were identified: 4 (1) 3.1 kg N/ha/ yr
– shift from sensitive to tolerant lichen species 5 (2) 5.2 kg N/ha/ yr
– dominance of the tolerant lichen species 6 (3) 10.2 kg N/ha/ yr
– loss of sensitive lichen sp ecies 7 (4) 17 kg N/ha/ yr
– leaching of nitrate into streams. 8 These benchmarks, ranging from 3.1 to 17 kg N/ha/y r, were compared to 2002 9 CMAQ/NADP data to discern an y associations between atmospheric deposition and changing 10 communities. Evidence supports the fi nding that nitrogen alters CSS and MCF. Key findings 11 include the following: 2002 CMAQ/NADP nitrogen deposition data show that the 3.3 kg N/ha/yr 12 benchmark has been exceeded in more than 93% of CSS areas (654,048 ha). These deposition 13 levels are a driving f orce in the degradation of CSS communities. Although CSS decline has 14 been observed in the absence of fire, the contributions of deposition and fire to the CSS decline 15 require further research. CSS is fragmented into many small parcels, and the 2002 16 CMAQ/NA DP 12
- km grid data are not fine enough to fully validate the relationship between 17 CSS distribution, nitrogen deposition, and fire. 2002 CMAQ/NADP nitrogen deposition data 18 exceeds the 3.1 kg N/ha/y r benchmark in more than 38% (1,099,133 ha) of MCF areas, an d 19 nitrate leaching has been observed in surface waters. Ozone effects confound nitrogen effects on 20 MCF acidoph yt e lichen, and the interrelationship between fire and nitrogen cycling requires 21 additional research. 22Draft

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59 c. Key uncertainties associated with nutrient e nrichment 1 There are different levels of uncertainty associated with relationships between deposition, 2 ecological effects and ecological indicators. The criteria used in the REA to evaluate the degree 3 of confidence in the data, modeling and ecological eff ect function are detailed in Chapter 7 of the 4 REA . Below is a discussion of uncertaint y relating aquatic and terrestrial ecosystems to nutrient 5 enrichment effects. 6 i. Aquatic ecosystems 7 The approach for assessing atmospheric contributions to total nitroge n loading in the 8 REA was to consider the main

- stem river to an estuary (including the estuary) rather than an 9 entire estuary

s ys tem or bay . The biological indicators used in the NOAA ASSETS EIrequired 10 the evaluation of man y national databases including the US Geological Survey National Water 11 Quality

Assessment (NAWQA) files, EPA’s STORage and RETrieval (STORET) database, 12 NOAA’s Estuarine Drainage Areas data, and EPA’s water quality standards nutrient criteria for 13 rivers and lakes ( US EPA, 200 9 , Appendix 6 and Table 1.2.

- 1 ). Both the SPARROW modeling 14 for nitrogen loads and assessment of estuary conditions under NOAA ASSETS EI , have been 15 applied on a national scale. The REA concludes that the available data are medium quality with 16 intermediate confidence a bout the use of these data and their values for extrapolating to a larger 17 regional area ( US EPA, 200 9 , section 7.3.1 ). Intermediate confidence is associated with the 18 modeling approach using ASSETS EIand SPARROW. The REA states there is low confidence 19 wi th the ecological effect function due to the results of the anal ys is which indicated that 20 reductions in atmospheric deposition alone could not solve coastal eutrophication problems due 21 to multiple non

- atmospheric nitrogen inputs ( US EPA, 200 9 , section 7.3. 3 ). 22Draft

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60 ii. Terrestrial ecosystems 1 Ecological thresholds are identified for CSS and MCF and these data are considered to be 2 of high quality , however, the ability to extrapolate these data to larger regional areas is limited 3 ( U S EPA, 2009 , section 7.4.1 ). No quantitative modeling was conducted or ecological effect 4 function developed for terrestrial nutrient enrichment reflecting the uncertainties associated with 5 these depositional effects. 6 4. Other Ecological Effects 7 It is stated in the ISA ( US EPA, 2008, section 3.4.1 and 4.5 ) that mercury is a highl y 8 neurotoxic contaminant that enters the food web as a methylated compound, methylmercury

. 9 Mercury

is principally

methylated by

sulfur

- reducing bacteria and can be taken up by 10 microorganisms, zooplankton and macroinvertebrates. The contaminant is concentrated in higher 11 trophic levels, including fish eaten b y humans. Experimental evidence has established that only 12 inconsequential amounts of methylmercury can be produced in the absence of sulfate. Once 13 methylmercury

is present, other variables influence how much accumulates in fish, but elevated 14 mercury

levels in fish can only

occur where substantial amounts of methylmercury are present. 15 Current evidence indicates that in watersheds where mercury is present, increased oxides of 16 sulfur deposition very likely results in additional production of methylmercury which leads to 17 greater accumulation of MeHg concentrations in fish. With respect to sulfur deposition and 18 mercury

methylation, the final I SA determined: The evidence is sufficient to infer a causal 19 relationship between sulfur deposition and increased mercury methylation in wetlands and 20 aquatic environments. 21 The production of meaningful amounts of meth yl mercury (MeHg) requires the presence 22 of SO 4 2
- and merc ury , and where mercury is present, increased availability of SO 4 2
- results in 23Draft

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61 increased production of MeHg. There is increasing evidence on the relationship between sulfur 1 deposition and increased methylation of mercury in aquatic environments; this effect occurs onl y 2 where other factors are present at levels within a range to allow methylation. The production of 3 methylmercury

requires the presence of sulfate and mercury , but the amount of methylmercury 4 produced varies with ox ygen content, temperature, pH, and supply of labile organic carbon ( US 5 EPA, 2008, section 3.4 ). In watersheds where changes in sulfate deposition did not produce an 6 effect, one or several of those interacting factors were not in the range required for meaningful 7 methylation to occur ( US EPA, 2008, section 3.4 ). Watersheds with conditions known to be 8 conducive to mercury methylation can be found in the northeastern United States and 9 southeastern Canada. 10 While the relationship between sulfur and methylmercury production was concluded to 11 be causal in the IS A, the REA concluded that there was insufficient evidence to quantify the 12 relationship between sulfur and meth yl mercury . Therefore onl y a qualitative assessment was 13 included in Chapter 6 of the REA
. The Policy Assessment was then unable to make a 14 determination as to the adequacy of the existing SO 2 standards in protecting against welfare 15 effects associated with increased mercury methylation. 16 B. Risk and Exposure Assessment 17 The risk and exposure assessment conducted for the current review was developed to 18 describe potential risk from current and future deposition of oxides of nitrogen and sulfur to 19 sensitive ecos ys tems. The case stud y anal ys es in the REA show that there is confidence that 20 known o r anticipated adverse ecological effects are occurring under current ambient loadings of 21 nitrogen and sulfur in sensitive ecosy stems across the United States. An overview of the 22 material covered in the REA , a summary of the key

findings from the air quali ty analy ses, 23Draft

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62 acidification and nutrient enrichment case studies, and general conclusions from evaluating 1 additional welfare effects, are presented below. 2 1. Overview of the Risk and Exposure Assessment 3 The REA evaluates the relationships between atmospheric concentrations, deposition, 4 biologicall y relevant exposures, targeted ecos ys tem effects, and ecos ys tem services. To evaluate 5 the nature and magnitude of ecos ys tem responses associated with adverse effects, the REA also 6 examines various way s to quantify the relationships between air quality indicators, deposition of 7 biologicall y available forms of nitrogen and sulfur, ecologicall y relevant indicators relating to 8 deposition, exposure and effects on sensitive receptors, and related effects resulting in chan ges in 9
ecos ys tem structure and services. The intent is to determine the exposure metrics that incorporate 10 the temporal considerations (i.e., biologically relevant timescales), pathways, and ecologicall y 11 relevant indicators necessary to maintain the functio ning of these ecos ys tems. To the extent 12 feasible, the REA evaluates the overall load to the sy stem for nitrogen and sulfur, as well as the 13 variability

in ecos ys tem responses to these pollutants as well as evaluating the contributions of 14 atmosphericall y dep osited nitrogen and sulfur relative to the combined atmospheric loadings of 15 both elements. Since oxidized nitrogen is the listed criteria pollutant (currently measured b y the 16 ambient air quality

indicator NO 2 ) for the atmospheric contribution to total nitr ogen, the REA 17 examines the contribution of nitrogen oxides to total reactive nitrogen in the atmosphere, relative 18 to the contributions of reduced forms of nitrogen (e.g., ammonia, ammonium), to ultimately 19 assess how a meaningful secondary National Ambient Air Quality Standards (NAAQS) might be 20 structured. 21 The REA focuses on ecosy stem welfare effects that result from the deposition of total 22 reactive nitrogen and sulfur. Because ecos ys tems are diverse in biota, climate, geochemistry , and 23Draft

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63 hydrology , response to pollutant exposures can vary greatl y between ecos ystems. In addition, 1 these diverse ecos ys tems are not distributed evenly across the United States. To target nitrogen 2 and sulfur acidification and nitrogen and sulfur enrichment, the REA addresses four m ain 3 targeted ecos ys tem effects on terrestrial and aquatic sy stems identified by the IS A ( US EPA, 4 2008 ): a quatic acidification due to nitrogen and sulfur ; t errestrial acidification due to nitrogen 5 and sulfur ; a quatic nutrient enrichment, including eutrophic ation ; and t errestrial nutrient 6 enrichment. 7 In addition to these four targeted ecos ys tem effects, the REA also qualitatively addresses 8 the influence of sulfur oxides deposition on methylmercury production; nitrous oxide (N 2 O) 9 effects on climate; nitrogen effects on primary productivity

and biogenic greenhouse gas fluxes; 10 and phy

totoxic effects on plants. 11 Because the targeted ecosy stem effects outlined above are not evenl y distributed across 12 the United States, the REA identified case studies for each targ eted effects based on ecos ys tems 13 identified as sensitive to nitrogen and/or sulfur deposition effects. Eight case study areas and 14 two supplemental study areas (Rocky Mountain National Park and Little Rock Lake, WI) are 15 summarized in the REA based on ecos y stem characteristics, indicators, and ecos ys tem service 16 information. Case studies selected for aquatic acidification effects were the Adirondack 17 Mountains and Shenandoah National Park. Kane Experimental Forest in Pennsylvania and 18 Hubbard Brook Experimental Forest in New Hampshire were selected as case studies for 19 terrestrial acidification. Aquatic nutrient enrichment case study locations were selected in the 20 Potomac River Basin upstream of Chesapeake Bay and the Neuse River Basin upstream of the 21 Pamlico Sound in North Carolina. The Coastal Sage Scrub Communities in southern California 22 and the Mixed Confer Forest Communities in the San Bernardino and Sierra Nevada Mountains 23Draft

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64 of California were selected as case studies for terrestrial nutrient enrichment. Two supplemental 1 areas were also chosen, one in Rocky Mountain National Park for terrestrial nutrient enrichment 2 and one in L ittle Rock Lake, Wisconsin for aquatic nutrient enrichment. 3 2. Key findings 4 In summary , based on case study anal ys es, the REA co ncludes that known or anticipated 5 adverse ecological effects are occurring under current conditions and further concludes that these 6 adverse effects continue into the future

. Key findings from the air qualit y analyses, acidification 7 and nutrient enrichment case studies, as well as general conclusions from evaluating additional 8 welfare effects, are summarized below. 9
a. Air quality analyses 10 The air qualit y analyses in the REA encompass the current emissions sources of nitrogen 11 and sulfur, as well as atmospheric concentrations, estimates of deposition of total nitrogen, 12 policy

- relevant background, and non

- ambient loadings of nitrogen and sulfur to ecos ys tems, both 13 nationwide an d in the case study areas. Spatial fields of deposition were created using wet 14 deposition measurements from the National Atmospheric Deposition Program (NADP) National 15 Trends Network and dry deposition predictions from the 2002 Community Multi

- Scale Air 16 Qu ality

(CMAQ) model simulation. 17 (1) Total reactive nitrogen deposition and sulfur deposition are much greater in the East 18 compared to most areas of the West. 19 (2) These regional differences in deposition correspond to the regional differences in oxides of 20 nitrogen and SO2 concentrations and emissions, which are also higher in the East. Oxides of 21 nitrogen emissions are much greater and generally more widespread than NH 3 emissions 22 nationwide; high NH 3 emissions tend to be more local (e.g., eastern North Carolina) or s ub

- 23Draft

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65 regional (e.g., the upper Midwest and Plains states). The relative amounts of oxidized versus 1 reduced nitrogen deposition are consistent with the relative amounts of oxides of nitrogen 2 and NH 3 emissions. Oxidized nitrogen deposition exceeds reduced nitr ogen deposition in 3 most of the case stud y areas; the major exception being the Neuse River/Neuse River Estuary 4 Case Study

Area. 5 (3) R educed nitrogen deposition exceeds oxidized nitrogen deposition in the vicinity of local 6 sources of NH 3 . 7 (4) There can be relativel y large spatial variations in both total reactive nitrogen deposition and 8 sulfur deposition within a case stud y area; this occurs particularl y in those areas that contain 9 or are near a high emissions source of oxides of nitrogen , NH 3 , and/or SO 2 . 10 (5) The seaso nal patterns in deposition differ between the case study areas. For the case stud y 11 areas in the East, the season with the greatest amounts of total reactive nitrogen deposition 12 correspond to the season with the greatest amounts of sulfur deposition. Deposi tion peaks in 13 spring in the Adirondack, Hubbard Brook Experimental Forest, and Kane Experimental 14 Forest case stud y areas, and it peaks in summer in the Potomac River/Potomac Estuary , 15 Shenandoah, and Neuse River/Neuse River Estuary case study

areas. For the case stud y areas 16 in the West, there is less consistency in the seasons with greatest total reactive nitrogen and 17 sulfur deposition in a given area. In general, both nitrogen and/or sulfur deposition peaks in 18 spring or summer. The exception to this is the Sierra Nevada Range portion of the Mixed 19 Conifer Forest Case Study Area, in which sulfur deposition is greatest in winter. 20 b. Deposition

- related aquatic acidification 21 The role of aquatic acidification in two eastern United States areas

— northeastern New 22 York’ s Adirondack area and the Shenandoah area in Virginia

— was anal yz ed in the REA to 23Draft

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66 assess surface water trends in SO 4 2
- and NO 3 - concentrations and acid neutralizing capacity (ANC) 1 levels and to affirm the understanding that reductions in deposition could influence the risk of 2 acidification. Monitoring data from the EPA

- administered Temporall y In tegrated Monitoring of 3 Ecosy stems (TIME)/Long

- Term Monitoring (LTM) programs and the Environmental Monitoring 4 and Assessment Program (EMAP) were assessed for the y ears 1990 to 2006, and past, present, 5 and future water quality levels were estimated using both steady

- state and dy namic 6 biogeochemical models. 7 Although wet deposition rates for SO 2 and oxides of nitrogen in the Adirondack Case 8 Study

Area have reduced since the mid

- 1990s, current concentrations in are still well above pre

9 acidification (1860) conditions. Model of Acidification of Groundwater in Catchments (MAGIC) 10 modeling predicts NO 3 - and SO 4 2
- are 17
- and 5

- fold higher today , respectively . The estimated 11
ave rage ANC for 44 lakes in the Adirondack Case Study Area is 62.1 microequivalents per liter 12 (μeq/L ) (± 15.7 μeq/L); 78 % of all monitored lakes in the Adirondack Case Study Area have a 13 current risk of Elevated, Severe , or Acute . Of the 78%, 31% experience episodic acidification, 14 and 18% are chronicall y acidic today

. 15 (1) Based on the stead y - state critical load model for the y ear 2002, 18%, 28%, 44%, and 58% of 16 169 modeled lakes received combined total sulfur and nitrogen deposition that exceeded their 17 critical l oad, with critical ANC limits of 0, 20, 50, and 100 μeq/Lrespectively

. 18 (2) Based on a deposition scenario that maintains current emission levels to 2020 and 2050, the 19 simulation forecast indicates no improvement in water qualit y in the Adirondack Case Stud y 20 Area. The percentage of lakes within the Elevated to Acute Concern classes remains the same 21 in 2020 and 2050. 22Draft

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67 (3) Since the mid

- 1990s, streams in the Shenandoah Case Study Area have shown slight declines 1 in NO 3 and SO 4 2
- concentrations in surface waters. ANC le vels increased from about 50 2 μeq/Lin the earl y 1990 to >75 μeq/Luntil 2002, when ANC levels declined back to 1991

– 3 1992 levels. Current concentrations are still above pre

- acidification (1860) conditions. 4 MAGI C modeling predicts surface water concentration s of NO 3 and SO 4 2
- are 10
- and 32

- 5 fold higher today , respectively . The estimated average ANC for 60 streams in the 6 Shenandoah Case Stud y Area is 57.9 μeq/L (± 4.5 μeq/L ). 55% of all monitored streams in 7 the Shenandoah Case Study Area have a current risk of Elevated, Severe , or Acute . Of the 8 55%, 18% experience episodic acidification, and 18% are chronically acidic today

. 9 (4) Based on the stead y - state critical load model for the y ear 2002, 52%, 72%, 85%, and 93% of 10 60 modeled streams received combined total sulf ur and nitrogen deposition that exceeded 11 their critical load, with critical ANC limits of 0, 20, 50, and 100 μeq/Lrespectively

. 12 (5) Based on a deposition scenario that maintains current emission levels to 2020 and 2050, the 13 simulation forecast indicates that a large number of streams still have Elevated to Acute 14 problems with acidity . In fact, from 2006 to 2050, the percentage of streams with Acute 15 Concern increases b y 5%, while the percentage of streams in Moderate Concern decreases 16 by

5%. 17
c. Deposition

- related terrestrial acidification 18 The role of terrestrial acidification was examined in the REA using a critical load 19 anal ys is for sugar maple and red spruce forests in the eastern United States by using the base 20 cation to aluminum (Bc/Al) ratio in acidified forest soils as an indicator to assess the impact of 21 nitrogen and sulfur deposition on tree health. These are the two most commonly s tudied species 22 in North America for impacts of acidification. At a Bc/Al ratio of 1.2, red spruce growth can be 23Draft

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68 reduced b y 20%. Sugar maple growth can be reduced b y 20% at a Bc/Al ratio of 0.6. Key 1 findings of the case stud y are summarized below. 2 (1) Case stu dy

results suggest that the health of at least a portion of the sugar maple and red 3 spruce growing in the United States may have been compromised with acidifying total 4 nitrogen and sulfur deposition in 2002

. 2002 CMAQ/NADP total nitrogen and sulfur 5 deposi tion levels exceeded three selected critical loads in 3% to 75% of all sugar maple plots 6 across 24 states. The three critical loads ranged from 107 to 6,008 eq/ha/y r for the Bc/Al 7 ratios of 0.6, 1.2, and 10.0 (increasing levels of tree protection). The 200 2 CMAQ/NADP 8 total nitrogen and sulfur deposition levels exceeded three selected critical loads in 3% to 36% 9 of all red spruce plots across 8 states. The three critical loads ranged from 180 to 4,278 10 eq/ha/y

r for the Bc/Al ratios of 0.6, 1 .2, and 10.0 (incr easing levels of tree protection). 11 (2) The Simple Mass Balance model assumptions made for base cation weathering (Bcw) and 12 forest soil ANC input parameters are the main sources of uncertainty since these parameters 13 are rarel y measured and require researchers t o use default values. Bcw contributed 49% to 14 the total variability in the critical load estimates, and forest soil ANC contributed 46% to the 15 total variability

. 16 (3) The pattern of case study results suggests that nitrogen and sulfur acidifying deposition in th e 17 sugar maple and red spruce forest areas studied were very close to, if not greater than, the 18 critical loads for those areas and both ecos ys tems are likel y to be sensitive to any future 19 changes in the levels of deposition. 20 d. Deposition

- related aquatic nutri ent enrichment 21 The role of nitrogen deposition in two main stem rivers feeding their respective estuaries 22 was anal yz ed in the REA to determine if decreases in deposition could influence the risk of 23Draft

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69 eutrophication as predicted using the Assessment of Estuarine Trophic Status eutrophication 1 index (ASSETS EI ) scoring s ys tem in tandem with SPAtially Referenced Regression on 2 Watershed Attributes (SPARROW) modeling. This modeling approach provides a transferrable, 3 intermediate

- level anal ys is of the linkages betw een atmospheric deposition and receiving waters, 4 while providing results on which conclusions could be drawn. Future application of the methods 5 to case stud y areas where atmospheric deposition plays a larger role in the nitrogen loading to an 6 estuary

will likely

provide more tangible results. A summary of findings follows: 7 (1) 2002 CMAQ/NADP results showed that an estimated 40,770,000 kg of total nitrogen was 8 deposited in the Potomac River watershed. SPARROW modeling predicted that 7,380,000 kg 9 N/y

r of the depo sited nitrogen reached the estuary (20% of the total load to the estuary). The 10 overall ASSETS EIfor the Potomac River and Potomac Estuary was Bad

. 11 (2) To improve the Potomac River and Potomac Estuary ASSETS EIscore from Bad to Poor , 12 there is a slim chance that a decrease of at least 78% in the 2002 total nitrogen atmospheric 13 deposition load to the watershed would be required. 14 (3) 2002 CMAQ/NADP results showed that an estimated 18,340,000 kg of total nitrogen was 15 deposited in the Neuse River watershed. SPARROW m odeling predicted that 1,150,000 kg 16 N/y

r of the deposited nitrogen reached the estuary (26% of the total load to the estuary). The 17 overall ASSETS EIfor the Neuse River/Neuse River Estuary was Bad

. 18 (4) It was found that the Neuse River/Neuse River Estuary ASSE TS EIscore could not be 19 improved from Bad to Poor with decreases onl y in the 2002 atmospheric deposition load to 20 the watershed. Additional reductions would be required from other nitrogen sources within 21 the watershed. 22Draft

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70 The small effect of decreasing atmos pheric deposition in the Neuse River watershed is 1 because the other nitrogen sources within the watershed are more influential than atmospheric 2 deposition to the total nitrogen loadings to the Neuse River Estuary as estimated with the 3 SPARROW model. A wate rbody ’s response to nutrient loading depends on the magnitude (e.g., 4 agricultural sources have a high influence in the Neuse), spatial distribution, and other 5 characteristics of the sources within the watershed. 6
e. Deposition

- related terrestrial nutrient enri chment 7 California C oastal S age S crub (CSS) and M ixed C onifer F orest (MCF) communities 8 were the focus of the Terrestrial Nutrient Enrichment Case Stud ies of the REA. Geographic 9 information sy stems (GIS) analy sis supported a qualitative review of past field research to 10 identify

ecological benchmarks associated with CSS and my corrhizal communities, as well as 11 MCF’s nutrient

- sensitive acidophy te lichen communities, fine

- root biomass in Ponderosa pine, 12 and leached nitrate in receiving waters. These benchmarks, ranging from 3.1 to 17 kg N/ha/ yr, 13 were compared to 2002 CMAQ/NADP data to discern an y associations between atmospheric 14 deposition and changing communities. Evidence supports the finding that nitrogen alters CSS 15 and MCF. Key

findings include the following: 16 (1) 2002 CMAQ/NADP nitrogen deposition data show that the 3.3 kg N/ha/ yrbenchmark has 17 been exceeded in more than 93% of CSS areas (654,048 ha). These deposition levels are a 18 driving force in the degradation of CSS communities. Although CSS decline has been 19
o bserved in the absence of fire, the contributions of deposition and fire to the CSS decline 20 require further research. CSS is fragmented into many small parcels, and the 2002 21 CMAQ/NADP 12

- km grid data are not fine enough to fully validate the relationship b etween 22 CSS distribution, nitrogen deposition, and fire. 23Draft

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71 (2) 2002 CMAQ/NADP nitrogen deposition data exceeds the 3.1 kg N/ha/ yr benchmark in more 1 than 38% (1,099,133 ha) of MCF areas, and nitrate leaching has been observed in surface 2 waters. Ozone effects confo und nitrogen effects on MCF acidoph yt e lichen, and the 3 interrelationship between fire and nitrogen cycling requires additional research. 4 f. Additional effects 5 Ecological effects have also been documented across the United States where elevated 6 nitrogen d eposition has been observed, including the eastern slope of the Rocky Mountains 7 where shifts in dominant algal species in alpine lakes have occurred where wet nitrogen 8 deposition was only

about 1.5 kg N/ha/ yr. High alpine terrestrial communities have a low 9 capacity

to sequester nitrogen deposition, and monitored deposition exceeding 3 to 4 kg N/ha/ yr 10 could lead to community

- level changes in plant species, lichens, and mycorrhizae. 11 Additional welfare effects are documented, but examined less extensively , in the REA. 12 These effects include qualitative discussions related to visibility and materials damage, such as 13 corrosion, erosion, and soiling of paint and buildings which are being addressed in the particulate 14 matter (PM) NAAQS review currently underway . A discussion of the causal relationship 15 between sulfur deposition (as sulfate, SO 4 2
- ) and increased mercury methylation in wetlands and 16 aquatic environments is also included in the REA. On this subject the REA concludes that 17 decreases in sulfate deposition will likely result in decreases in methyl mercury concentration; 18 however, spatial and biogeochemical variations nationally hinder establishing large scale dose

19 response relationshi ps. 20 Several additional issues concerning oxides of nitrogen were addressed in the REA. 21 Consideration was also given to nitrous oxide (N 2 O), a potent GHG. The REA concluded that it 22 is most appropriate to analyze the role of N 2 O in the context of all of the GHGs rather than as 23Draft

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72 part of the REA for this review. The REA considered n itrogen deposition and its correlation 1 with the rate of photos yn thesis and net primary productivity

. Nitrogen addition ranging from 2 15.4 to 300 kg N/ha/ yr is documented as incre asing wetland N 2 O production by an average of 3 207%. Nitrogen addition ranging from 30 to 240 kg N/ha/ y increased CH 4 emissions by 115%, 4 averaged across all ecosystems, and methane uptake was reduced b y 38% when nitrogen addition 5 ranged from 10 to 560 kg N/ha/y r, but reductions were onl y significant for coniferous and 6 deciduous forests. The heterogeneit y of ecos ys tems across the United States, however, 7 introduces variations into dose

- response relationships. 8 The p hy

totoxic effects of oxides of nitrogen and s ulfur on vegetation were also briefl y 9 discussed in the REA which concluded that since a unique secondary NAAQS exists for SO 2 , 10 and concentrations of nitric oxide (NO ) , NO 2 , and PAN are rarely high enough to have 11 phy totoxic effects on vegetation that furthe r assessment was not warranted at this time. 12 3. Conclusions on E ffects 13 For aquatic and terrestrial acidification effects, a similar conceptual approach was used 14 (critical loads) to evaluate the impacts of multiple pollutants on an ecological e ndpoint, wher eas 15 the approaches used for aquatic and terrestrial nutrient enrichment were fundamentally distinct. 16 Although the ecological indicators for aquatic and terrestrial acidification (i.e., ANC and Bc/Al) 17 are very

different, both ecological indicators are well

- correlated with effects such as reduced 18 biodiversity

and growth. While aquatic acidification is clearl y the targeted effect area with the 19 highest level of confidence, the relationship between atmospheric deposition and an ecological 20 indicator is also quite strong for terrestrial acidification. The main drawback with the 21 understanding of terrestrial acidification is that the data are based on laboratory responses rather 22Draft

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73 than field measurements. Other stressors that are present in the field but that are not p resent in 1 the laboratory

may confound this relationship. 2 For nutrient enrichment effects, the REA utilized different t ypes of indicators for aquatic 3 and terrestrial effects to assess both the likelihood of adverse effects to ecosy stems and the 4 relationshi p between adverse effects and atmospheric sources of oxides of nitrogen. The 5 ecological indicator chosen for aquatic nutrient enrichment, the ASSETS EI, seems to be 6 inadequate to relate atmospheric deposition to the targeted ecological effect, likely due to the 7 many

other confounding factors. Further, there is far less confidence associated with the 8 understanding of aquatic nutrient enrichment because of the large contributions from non

9 atmospheric sources of nitrogen and the influence of both oxidized and reduced forms of 10 nitrogen, particularl y in large watersheds and coastal areas. However, a strong relationship exists 11 between atmospheric deposition of nitrogen and ecological effects in high alpine lakes in the 12 Rocky

Mountains because atmospheric depositi on is the only source of nitrogen to these 13 sy stems. There is also a strong weight

- of

- evidence regarding the relationships between 14 ecological effects attributable to terrestrial nitrogen nutrient enrichment; however, ozone and 15 climate change may be confound ing factors. In addition, the response for other species or species 16 in other regions of the United States has not been quantified. 17 C. Adversity of Effects to Public Welfare 18 Characterizing a known or anticipated adverse effect to public welfare is an important 19 component of developing any secondary

NAAQS. According to the Clean Air Act, welfare 20 effects include: “ effects on soils, water, crops, vegetation, manmade materials, animals, wildlife, 21 weather, visibility , and climate, damage to and deterioration of prop ert y, and hazards to 22 transportation, as well as effect on economic values and on personal comfort and well

- being, 23Draft

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74 whether caused b y transformation, conversion, or combination with other air pollutants ” (CAA, 1 Section 302(h)). While the text above lists a number of welfare effects, these effects do not 2 define public welfare in and of themselves. 3 Although there is no specific definition of adversity to public welfare, the paradigm of 4 linking adversit y to public welfare to disruptions in ecos ys tem structure and function has been 5 used broadl y b y EPA to categorize effects of pollutants from the cellular to the ecos ys tem level. 6 An evaluation of adversity to public welfare might consider the likelihood, ty pe, magnitude, and 7 spatial scale of the effect as well as the potential for recovery and any uncertainties relating to 8 these considerations. 9 Similar concepts were used in past reviews of secondary NAAQS for ozone, PM (relating 10 to visibility), as well as in initial reviews of effects from lead deposition. Beca use oxides of 11 nitrogen and sulfur are deposited from ambient sources into ecos ys tems where they affect 12 changes to organisms, populations and ecos ys tems, the concept of adversity to public welfare as 13 related to impacts on the public from alterations in stru cture and function of ecos ys tems is an 14 appropriate consideration for this review. 15 Based on information provided in the Policy Assessment , the following section discusses 16 how ecological effects from deposition of oxides of nitrogen and sulfur relate to adversit y to 17 public welfare

. In the Policy Assessment , p ublic welfare was discussed in terms of loss of 18
ecos ys tem services (defined below) , which in some cases can be monetized. Each of the four 19 main effect areas (aquatic and terrestrial acidification and aquatic and terrestrial nutrient over

20 enrichment) are discussed including current ecological effects and associated ecos ys tem services, 21 including potential value of the services. 22Draft

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75 1. Ecosystem Services 1 The Policy

Assessment defines ecosy stem services as the benefits individuals and 2 organizations obtain from ecosy stems. Ecos ys tem services can be classified as provisioning 3 (food and water), regulating (control of c limate and disease), cultural (recreational, existence, 4 spiritual, educational), and supporting (nutrient cycling). Conceptuall y, changes in ecos ys tem 5 services may

be used to aid in characterizing a known or anticipated adverse effect to public 6 welfare. In the REA and Policy Assessment ecosy stem services are discussed as a method of 7 assessing the magnitude and significance to the public of resource s affected by ambient 8 concentrations of oxides of nitrogen and sulfur and deposition in sensitive ecos ys tems
. 9 As EPA has in previous NAAQS reviews defined ecological goods and services for the 10 purposes of a Regulatory Impact Anal ys is as the “outputs of ecological functions or processes 11 that directly

or indirectl y contribute to social welfare or have the potenti al to do so in the future. 12 Some outputs may

be bought and sold, but most are not marketed.” It is e specially important to 13 acknowledge that it is difficult to measure and/or monetize the goods and services supplied by 14
ecos ys tems. It can be informative i n characterizing adversit y to public welfare to attempt to 15 place an economic valuation on the set of goods and services that have been identified with 16 respect to a change in policy however it must be noted that this valuation will be incomplete and 17 illustr ative only

. 18 Knowledge about the relationships linking ambient concentrations and ecosy stem 19 services is considered in the Policy Assessment as one method by which to inform a policy 20 judgment on a known or anticipated adverse public welfare effect. For ex ample, a change in an 21
ecos ys tem structure and process, such as foliar injury , would be classified as an ecological effect, 22 with the associated changes in ecos ys tem services, such as primary productivity , food 23Draft

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76 availability , forest products, and aesthetics ( e.g., scenic viewing), classified as public welfare 1 effects. Additionally , changes in biodiversity would be classified as an ecological effect, and the 2 associated changes in ecosy stem services

— productivity , existence (nonuse) value, recreational 3 viewing and aesthetics

—would also be classified as public welfare effects. 4 As described in Chapters 4 and 5 of the REA , case study analy ses were performed that 5 link deposition in sensitive ecos ys tems to changes in a given ecological indicator (e.g., for 6 aquatic acidification, to changes in acid neutralizing capacity [ANC]) and then to changes in 7
ecos ys tems. Appendix 8 of the REA links the changes in ecos ys tems to the services they 8 provide (e.g., fish species richness and its influence on recreational fishing). T o the extent 9 possible for each targeted effect area, the REA linked ambient concentrations of nitrogen and 10 sulfur (i.e., ambient air quality indicators) to deposition in sensitive ecosy stems (i.e., exposure 11 pathway s), and then to system response as measure d by a given ecological indicator (e.g., lake 12 and stream acidification as measured b y ANC). The ecological effect (e.g., changes in fish 13 species richness) was then, where possible, associated with changes in ecosy stem services and 14 the corresponding public welfare effects (e.g., recreational fishing). 15 2. Effects on Ecosystem Services 16 The process used to link ecological indicators to ecosy stem services is discussed 17 extensively

in Appendix 8 of the REA. In brief, for each case study area assessed the ecological 18 indicators are linked to an ecological response that is subsequently linked to associated services 19 to the extent possible. For example in the case study for aquatic acidification the chosen 20 ecological indicator is ANC which can be linked to the ecos ys tem service of recreational fishing. 21 Although recreational fishing losses are the onl y s ervice effects that can be independentl y 22Draft

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77 quantified or monetized at this time, there are numerous other ecos ys tem services that may be 1 related to the ecological effects of acidification. 2 While aquatic acidification is the focus of this proposed standard, the other effect areas 3 were also anal yz ed in the REA and these ecos ys tems are being harmed by nitrogen and sulfur 4 deposition and will obtain some measure of pro tection with any decrease in that deposition 5 regardless of the reason for the decrease. The following summarizes the current levels of specific 6
ecos ys tem services for aquatic and terrestrial acidification, and aquatic and terrestrial nutrient 7
over - enrichme nt and attempts to quantify and when possible monetize the harm to public 8 welfare, as represented by ecos ys tem services, due to nitrogen and sulfur deposition. 9
a. Aquatic Acidification 10 Acidification of aquatic ecosy stems primarily affects the ecos ys tem serv ices that are 11 derived from the fish and other aquatic life found in surface waters. In the northeastern United 12 States, the surface waters affected b y acidification are not a major source of commerciall y raised 13 or caught fish; however, they are a source of food for some recreational and subsistence fishers 14 and for other consumers. Although data and models are available for examining the effects on 15 recreational fishing, relatively little data are available for measuring the effects on subsistence 16 and other c onsumers. Inland waters also provide aesthetic and educational services along with 17
non - use services, such as existence value (protection and preservation with no expectation of 18 direct use). In general, inland surface waters such as lakes, rivers, and str eams also provide a 19 number of regulating services, playing a role in hydrological regimes and climate regulation. 20 There is little evidence that acidification of freshwaters in the northeastern United States has 21 significantl y degraded these specific service s; however, freshwater ecos ys tems also provide 22 biological control services by providing environments that sustain delicate aquatic food chains. 23Draft

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78 The toxic effects of acidification on fish and other aquatic life impair these services b y disrupting 1 the troph ic structure of surface waters. Although it is difficult to quantify these services and how 2 they are affected by

acidification, it is worth noting that some of these services may be captured 3 through measures of provisioning and cultural services. For example, these biological control 4 services may

serve as “intermediate” inputs that support the production of “final” recreational 5 fishing and other cultural services. 6 As summarized in Chapter 4 of the Policy Assessment , recent studies indicate that 7 acidification of lakes and streams can result in significant loss in economic value. For example, 8 data indicate that more than 9% of adults in the northeastern part of the country participate 9 annually

in freshwater fishing yielding 140 million freshwater fishing day s. Each fishing day 10 has an estimated average value per day of $35. Therefore, the implied total annual value of 11 freshwater fishing in the northeastern United States was $5 billion in 2006. Embedded in these 12 numbers is a degree of harm to recre ational fishing services due to acidification that has occurred 13 over time. These harms have not been quantified on a regional scale; however, a case study was 14 conducted in the Adirondacks area ( US EPA, 20 11 , section 4.4.2). 15 In the Adirondacks case study, estimates of changes in recreational fishing services were 16 determined, as well as changes more broadl y in “cultural” ecos ys tem services (including 17 recreational, aesthetic, and nonuse services). First, the MAGIC model (U S EPA, 2009, 18 Appendix 8 and s ec ti on 2.2) was applied to 44 lakes to predict what ANC levels would be under 19 both “business as usual” conditions (i.e., allowing for some decline in deposition due to existing 20 regulations) and pre

- emission (i.e., background) conditions. Second, to estimate the recreational 21 fishing impacts of aquatic acidification in these lakes, an existing model of recreational fishing 22 demand and site choice was applied. This model predicts how recreational fishing patterns in the 23Draft

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79 Adirondacks would differ and how much high er the average annual value of recreational fishing 1 services would be for New York residents if lake ANC levels corresponded to background 2 (rather than business as usual) conditions. To estimate impacts on a broader category of cultural 3 (and some provisio ning) ecos ys tem services, results from the Banzhaf et al (2006) valuation 4 survey

of New York residents were adapted and applied to this context. The survey used a 5 contingent valuation approach to estimate the average annual household WTP for future 6 reduc tions in the percent of Adirondack lakes impaired by acidification. The focus of the survey 7 was on impacts on aquatic resources. Pretesting of the survey indicated that respondents 8 nonetheless tended to assume that benefits would occur in the condition of birds and forests as 9 well as in recreational fishing. 10 By

extrapolating the 44 lake Adiriondack case study to all 3,000 Adirondack lakes and b y 11 applying the willingness to pay survey results to all New York residents, the study estimated 12 aggregated benef

its between $300 and $800 million annually for the equivalent of improving 13 lakes in the Adirondacks region to an ANC level of 50 μeq/L . The REA estimated 44% of the 14 Adirondack lakes currently fall below an ANC of 50 μeq/L . Several states have set goals f or 15 improving the acid status of lakes and streams, generall y targeting ANC in the range of 50 to 60 16 μeq/L , and have engaged in costly activities to decrease acidification. 17 These results imply significant value to the public in addition to those derived from 18 recreational fishing services. Note that the results are onl y applicable to improvements in the 19 Adirondacks valued b y residents of New York. If similar benefits exist in other acid

- impacted 20 areas, benefits for the nation as a whole could be substant ial. The anal ys is provides results on 21 only

a subset of the impacts of acidification on ecosy stem services and suggests that the overall 22 impact on these services is likely to be substantial. 23Draft

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80 b. Terrestrial Acidification 1 Chapter 4.4.3 and 4.4.4 of the Policy A ssessment review several economic studies of 2 areas sensitive to terrestrial acidification. Forests in the northeastern United States provide 3 several important and valuable provisioning ecos ystem services, which are reflected in the 4 production and sales of tree products. Sugar maples are a particularly important commercial 5 hardwood tree species in the United States, producing timber and maple s yrup that provide 6 hundreds of millions of dollars in economic value annually . Red spruce is also used in a variet y 7 of wood products and provides up to $100 million in economic value annually . Although the 8 data do not exist to directly link acidification damages to economic values of lost recreational 9
ecos ys tem services in forests, these resources are valuable to the public. A recent stud y, 10 reviewed in the Policy Assessment, suggests that the total annual value of recreational off

- road 11 driving was more than $9 billion and the value of hunting and wildlife viewing was more than $4 12 billion each in the northeastern Unit ed States. 13
c. Nutrient Enrichment 14 Chapter 4.4.5 and 4.4.6 of the Policy Assessment summarize economic studies of east 15 coast estuaries affected by nutrient over

- enrichment or eutrophication. Estuaries in the eastern 16 United States are important for fish and shellfish production. The estuaries are capable of 17 supporting large stocks of resident commercial species, and they serve as the breeding grounds 18 and interim habitat for several migratory species. To provide an indication of the magnitude of 19 provisioning s ervices associated with coastal fisheries, from 2005 to 2007, the average value of 20 total catch was $1.5 billion per year in 15 East Coast states. Estuaries also provide an important 21 and substantial variet y of cultural ecos ys tem services, including water

- b ased recreational and 22 aesthetic services. For example, data indicate that 4.8% of the population in coastal states from 23Draft

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81 North Carolina to Massachusetts participated in saltwater fishing, with a total of 26 million 1 saltwater fishing day s in 2006. Based on estimates in the Policy Assessment , total recreational 2 value from these saltwater fishing day s was approximately $1.3 billion. Recreational 3 participation estimates for 1999

–2000 showed almost 6 million individuals participated in 4 motorboating in coastal s tates from North Carolina to Massachusetts. The aggregate value of 5 these coastal motorboating outings was $2 billion per year. 6 Terrestrial ecos ys tems can also suffer from nutrient over

- enrichment. Each ecos ys tem is 7 different in its composition of species and nutrient requirements. Changes to individual 8
ecos ys tems from changes in nitrogen deposition can be hard to assess economically . Relative 9 recreational values are often determined b y public use information. Chapter 4.4.7 of the Policy 10 Assessment revi ewed studies related to park use in California. Data from California State Parks 11 indicate that in 2002, 68.7% of adult residents participated in trail hiking for an average of 24.1 12 day s per y ear. The anal ys es in the Policy Assessment indicate that the agg regate annual benefit 13 for California residents from trail hiking in 2007 was $11.59 billion. 14 The Policy

Assessment also identified fire regulation as a service that could be affected 15 by nutrient over

- enrichment of the Coastal Sage Scrub and Mixed Conifer Forest ecos ys tems by 16 encouraging growth of more flammable grasses, increasing fuel loads, and altering the fire cycle. 17 Over the 5

- ye ar period from 2004 to 2008, Southern California experienced, on average, over 18 4,000 fires per ye ar, burning, on average, ov er 400,000 acres per ye ar. It is not possible at this 19 time to quantify

the contribution of nitrogen deposition, among man y other factors, to increased 20 fire risk. 21Draft

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82 3. Summary 1 Adversity

to public welfare can be understood not only by looking at how deposition of 2 oxides of nitrogen and sulfur affect the ecological functions of an ecos ys tem (see II.A.), but also 3 by understanding the ecosy stem services that are degraded. The monetized value of the 4
ecos ys tem services provided by ecos ys tems that are sensitive to de position of oxides of nitrogen 5 and sulfur are in the billions of dollars each year. Many lakes and streams are known to be 6 degraded b y acidic deposition which affects recreational fishing and tourism. Forest growth is 7 likely

suffering from acidic deposit ion in sensitive areas affecting red spruce and sugar maple 8 timber production, sugar maple sy rup production, hiking, aesthetic enjoy ment and tourism. 9 Nitrogen deposition contributes significantly to eutrophication in many estuaries affecting fish 10 producti on, swimming, boating, aesthetic enjo ym ent, and tourism. Important biodiversity is 11 likely

affected in many

natural and scenic terrestrial areas affecting habitat for rare and 12 endangered species, fire control, hiking, aesthetic enjoy ment, and tourism. 13 D. Adequacy of the Current Standards 14 An important issue to be addressed in the current review of the secondary standards for 15 oxides and nitrogen and sulfur standard is whether, in view of the scientific evidence reflected in 16 the I SA, additional information on exposure and risk discussed in the REA, and conclusions 17 drawn from the Policy Assessment , the existing standards provide adequate protection. The 18 Administrator therefore, has considered the extent to which the current standards are adequate 19 for the pr otection of public welfare. Having reached this general conclusion, it is then 20 necessary

to first evaluate the appropriateness of the current standards to address the ecological 21 effects of oxides of nitrogen and sulfur as well as the adequacy of the cur

rent secondary 22 standards for oxides of nitrogen and sulfur to provide requisite protection by considering to what 23Draft

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83 degree risks to sensitive ecos ys tems would be expected to occur in areas that meet the current 1 standards. Conclusions regarding the adequacy of the current standards are based on the 2 available ecological effects, exposure and risk

- based evidence. In evaluating the strength of this 3 information, EPA has taken into account the uncertainties and limitations in the scientific 4 evidence. This secti on addresses the adequacy of the current standards to protect against direct 5 exposure effects on plants from oxides of nitrogen and sulfur, the appropriateness of the current 6 structure of the standards to address deposition

- related effects of oxides of nitrogen and sulfur on 7 sensitive ecos ys tems and finally , the adequacy of such standards to protect against adverse 8 effects related to the deposition of oxides of nitrogen and sulfur. 9 1. Adequacy of the Current Standards for Direct Effects 10 The current secon dary oxides of nitrogen and sulfur standards are intended to protect 11 against adverse effects to public welfare. For oxides of nitrogen, the current secondary standard 12 was set identical to the primary standard 3 , e.g. an annual standard set for NO 2 to prote ct against 13 adverse effects on vegetation from direct exposure to ambient oxides of nitrogen. For oxides of 14 sulfur, the current secondary standard is a 3

- hour standard intended to provide protection for 15 plants from the direct foliar damage associated with atmospheric concentrations of SO 2 . It is 16 appropriate to consider whether the current standards are adequate to protect against the direct 17 effects on vegetation resulting from ambient NO 2 and SO 2 which were the basis for the current 18 secondary

standards. T he I SA concluded that there was sufficient evidence to infer a causal 19 relationship between exposure to SO 2 , NO, NO 2 and PAN and injury to vegetation. Additional 20 research on acute foliar injury has been limited and there is no evidence to suggest foliar in jury 21 below the levels of the current secondary standards for oxides of nitrogen and sulfur. There is 22

3 The current primary NO 2 standard has recently been changed to the 3 year average of the 98 th percentile of the annual distribution of the 1 hour daily maximum of the concentration of NO 2 . The current secondary standard rem ains as it was set in 1971.Draft

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84 sufficient evidence to suggest that the levels of the current standards are likely adequate to 1 protect against direct ph ytotoxic effects. 2 2. Appropriateness and Adequacy of the Current Standards for Deposition

- related Effects 3 This section addresses two concepts necessary to evaluate the current standards on the 4 context of deposition related effects. First, appropriateness of the current stand ards is considered 5 with regard to indicator, form, level and averaging time. This discussion centers around the 6 ability

of the current standards to evaluate and provide protection against deposition related 7 effects that vary

spatiall y and temporall y. It includes particular emphasis on the indicators and 8 forms of the current standards and the degree to which they are ecologically

relevant with regard 9 to deposition related effects. Second, this section relates the current standards and adequacy of 10 protectio n. 11
a. Appropriateness 12 The I SA has established that the major effects of concern for this review of the oxides of 13 nitrogen and sulfur standards are associated with deposition of N and S caused by atmospheric 14 concentrations of oxides of nitrogen and sulfur. The current standards are not directed toward 15 depositional effects, and none of the elements of the current NAAQS

– indicator, form, 16 averaging time, and level

– are suited for addressing the effects of N and S deposition. 17 Four issues arise that call i nto question the ecological relevance of the structure of the 18 current secondary

standards for oxides of nitrogen and sulfur. 19 (1) The current SO 2 secondary standard (0.5 ppm SO 2 over a 3
- hour average) does not utilize an 20 exposure period that is relevant for ecos ys tem impacts. The majorit y of deposition related 21 impacts are associated with depositional loads that occur over periods of months to years. 22 This differs significantl y from exposures associated with hourly concentrations of SO 2 as 23Draft

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85 measured b y the cur

rent secondary standard. By addressing short

- term concentrations, the 1 current SO 2 secondary standard, while protective against direct foliar effects from gaseous 2 oxides of sulfur, does not take into account the findings of effects in the I SA, which notes the 3 relationship between annual deposition of S and acidification effects which are likel y to be 4 more severe and widespread than ph yt otoxic effects under current ambient conditions, and 5 include effects from long term deposition as well as short term. Acid ification is a process 6 which occurs over time, as the ability of an aquatic sy stem to counteract acidic inputs is 7 reduced as natural buffers are used more rapidly than they can be replaced through geologic 8 weathering. The relevant period of exposure for ec osy stems is therefore not the exposures 9 captured in the short averaging time of the current SO 2 secondary standard. The current 10 secondary

standard for oxides of nitrogen is an annual standard (0.053 ppm averaged over 1 11 ye ar) and as such is more ecologicall y relevant. Current standards do not utilize appropriate 12 atmospheric indicators. NO 2 and SO 2 are used as the component of oxides of nitrogen and 13 sulfur that are measured, but they do not provide a complete link to the direct effects on 14
ecos ys tems from dep osition of oxides of nitrogen and sulfur as they do not capture all 15 relevant chemical species of oxidized nitrogen and oxidized sulfur that contribute to 16 deposition. The I SA provides evidence that deposition related effects are linked with total 17 nitrogen and total sulfur deposition, and thus all forms of oxidized nitrogen and oxidized 18 sulfur that are deposited will contribute to effects on ecos ys tems. Thus, by using 19 atmospheric NO 2 and SO 2 concentrations as indicators, the current standards address onl y a 20 fraction of total atmospheric oxides of nitrogen and sulfur, and do not take into account the 21 effects from deposition of total atmospheric oxides of nitrogen and sulfur. This suggests that 22Draft

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86 more comprehensive atmospheric indicators should be considered in designing ecologicall y 1 relevant standards. 2 (2) Current standards reflect separate assessments of the two individual pollutants, NO 2 and SO 2 , 3 rather than assessing the joint impacts of deposition to ecos ys tems. Recognizing the role that 4 each pollutant play s in jointly affecting ecos ys tem indicators, functions, and services is vital 5 to developing a meaningful standard. The clearest example of this interaction is in 6 assessment of the impacts of acidifying deposition on aquatic ecosy stems. Acidification in 7
an aquatic ecos ys tem depends on the total acidifying potential of the deposition of both N 8 and S from both atmospheric deposition of oxides of nitrogen and sulfur as well as the inputs 9 from other sources of N and S such as reduced nitrogen and non

- atmospheri c sources. It is 10 the joint impact of the two pollutants that determines the ultimate effect on organisms within 11 the ecos ys tem, and critical ecos ys tem functions such as habitat provision and biodiversity

. 12 Standards that are set independentl y are less able to account for the contribution of the other 13 pollutant. This suggests that interactions between oxides of nitrogen and oxides of sulfur 14 should be a critical element of the conceptual framework for ecologically relevant standards. 15 There are also important interactions between oxides of nitrogen and sulfur and reduced 16 forms of nitrogen, which also contribute to acidification and nutrient enrichment. Although 17 the standards do not directly address reduced forms of nitrogen in the atmosphere, e.g. they 18 do not require specific levels of reduced nitrogen, it is important that the structure of the 19 standards address the role of reduced nitrogen in determining the ecological effects resulting 20 from deposition of atmospheric oxides of nitrogen and sulfur. Considerat ion will also have 21 to be given to total loadings as ecos ys tems respond to all sources of N and S. 22Draft

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87 (3) Current standards do not take into account variability in ecos ys tem sensitivity . Ecos ys tems 1 are not uniforml y distributed either spatiall y or temporally in their sensitivity to air pollution. 2 Therefore, failure to account for the major determinants of variability , including geological 3 and soil characteristics related to the sensitivity to acidification or nutrient enrichment as well 4 as atmospheric and landscap e characteristics that govern rates of deposition, may lead to 5 standards that do not provide requisite levels of protection across ecos ys tems. The current 6 structures of the standards do not address the complexities in the responses of ecos ys tems to 7 deposi tion of oxides of nitrogen and sulfur. Ecos ystems contain complex groupings of 8 organisms that respond in various way s to the alterations of soil and water that result from 9 deposition of nitrogen and sulfur compounds. Different ecos ys tems therefore respon d in 10 different way s depending on a multitude of factors that control how deposition is integrated 11 into the sy stem. For example, the same levels of deposition falling on limestone dominated 12 soils have a very different effect than those falling on shallow g laciated soils underlain with 13 granite. One s ys tem may over time display no obvious detriment while the other may 14 experience a catastrophic loss in fish communities. This degree of sensitivity is a function 15 of many

atmospheric factors which control rates of deposition as well as ecological factors 16 which control how an ecosy stem responds to that deposition. The current standards do not 17 take into account spatial and seasonal variations not only in depositional loadings but also in 18 sensitivity

of ecos ys tems exposed to those loadings. Based on the discussed summarized 19 above, the Policy Assessment concludes that the current secondary standards for oxides of 20 nitrogen and oxide sof sulfur are not ecologicall y relevant in terms of averaging time, form, 21 level or i ndicator. 22Draft

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88 b. Adequacy of Protection 1 As described in the Policy Assessment , 2005 ambient conditions indicate that the current 2 SO 2 and NO 2 secondary standards are not exceeded ( US EPA, 20 11 , Figures 6

- 1 and 6

- 2) in 3 locations where ecological effects have been observed. In many locations, SO 2 and NO 2 4 concentrations are substantially below the levels of the secondary standards. This pattern 5 suggests that levels of deposition and any effects on ecos ys tems due to deposition of oxides of 6 nitrogen and sulfur under recent conditions are occurring even though areas meet or are below 7 current standards. In addition, these lev els based on conclusions in the REA will not decline in 8 the future to levels below which it is reasonable to anticipate effects. 9 Indetermining the adequacy of the current secondary standards for oxides of nitrogen 10 and sulfur the Policy Assessment consid er ed the extent to which ambient deposition contributes 11 to loadings in ecos ys tems. Since the last review of the secondary standard for oxides of nitrogen, 12 a great deal of information on the contribution of atmospheric deposition associated with 13 ambient ox ides of nitrogen has become available. The REA presents a thorough assessment of 14 the contribution of oxidized nitrogen to nitrogen deposition throughout the U.S., and the relative 15 contributions of ambient oxidized and reduced forms of nitrogen. The REA c oncludes that based 16 on that anal ys is, ambient oxides of nitrogen are a significant component of atmospheric nitrogen 17 deposition, even in areas with relatively high rates of deposition of reduced nitrogen. In 18 addition, atmospheric deposition of oxidized ni trogen contributes significantly to total nitrogen 19 loadings in nitrogen sensitive ecosy stems. 20 The I SA summarizes the available studies of relative nitrogen contribution and finds that 21 in much of the U.S., oxides of nitrogen contribute from 50 to 75 perce nt of total atmospheric 22 deposition relative to total reactive nitrogen that includes oxidized and reduced nitrogen species 23Draft

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89 ( US EPA, 2008, section 2.8.4 ). Although the proportion of total nitrogen loadings associated 1 with atmospheric deposition of nitrogen varies across locations (N deposition in the eastern U.S. 2 includes locations with greater than 9 kg N/ha

- yr , and in the central U.S. high deposition 3 locations with values on the order of 6 to 7 kg N/ha

- yr), the IS A indicates that atmospheric N 4 deposition is the main source of new anthropogenic N to most headwater streams, high elevation 5 lakes, and low

- order streams. Atmospheric N deposition contributes to the total N load in 6 terrestrial, wetland, freshwater, and estuarine ecosy stems that receive N through multiple 7 pathway s. In several large estuarine s ys tems, including the Chesapeake Bay , atmospheric 8 deposition accounts for between 10 and 40 percent of total nitrogen loadings (U S EPA, 2008 ). 9 Atmospheric concentrations of oxides of sulfur account for nearl y all S deposition in the 10 US. For the period 2004

–2006, mean S deposition in the U.S. was greatest east of the 11 Mississippi River with the highest deposition amount, 21.3 kg S/ha

- yr, in the Ohio River Valley 12 where most recording stations reported 3 year averages >10 kg S/ha

- yr . Numerous other stations 13 in the East reported S deposition >5 kg S/ha

- yr. Total S deposition in the U.S. west of the 100th 14 meridian was relativel y low, with all recording stations reporting <2 kg S/ha

- yr and man y 15 reporting <1 kg S/h a - yr. S was primarily deposited in the form of wet SO 4 2

− followed in 16 decreasing order b y a smaller proportion of dry SO 2 and a much smaller proportion of deposition 17 as dry

SO 4 2

−
. 18 As discussed throughout the REA ( US EPA, 200 9 and s ection IIB above), there are 19 several key

areas of risk that are associated with ambient concentrations of oxides of nitrogen 20 and sulfur. As noted earlier, in previous reviews of the secondary standards for oxides of 21 nitrogen and sulfur, the standards were designed to protect against direct exposure of plants to 22 ambient concentrations of the pollutants. A significant shift in understanding of the effects of 23Draft

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90 oxides of nitrogen and sulfur has occurred since the last reviews, reflecting the large amount of 1 research that has been conducted on the effects of deposition of nitrogen and sulfur to 2
ecos ys tems. The most significant risks of adverse effects to public welfare are those related to 3 deposition of oxides of nitrogen and sulfur to both terrestrial and aquatic ecosy stems. These risks 4 fall into two categories, acidification and nutrient enrichment, which were emphasized in the 5 REA are most relevant to evaluating the adequacy of the existing standards in protecting public 6 welfare from adverse ecological effects. 7 i. Aquatic acidificat ion 8 The focus of the REA case studies was on determining whether deposition of sulfur and 9 oxidized nitrogen in locations where ambient oxides of nitrogen and sulfur were at or below the 10 current standards was resulting in acidification and related effects , including episodic 11 acidification and mercury meth yl ation. Based on the case studies conducted for lakes in the 12 Adirondacks and streams in Shenandoah National Park (case studies are discussed more full y in 13
s ection IIB and US EPA, 2009 ), there is signific ant risk to acid sensitive aquatic ecos ys tems at 14 atmospheric concentrations of oxides of nitrogen and sulfur at or below the current standards. 15 The REA also supports strongl y a relationship between atmospheric deposition of oxides of 16 nitrogen and sulfur and loss of ANC in sensitive ecosy stems and that ANC is an excellent 17 indicator of aquatic acidification. The REA also concludes that at levels of deposition associated 18 with oxides of nitrogen and sulfur concentrations at or below the current standards, ANC levels 19 are expected to be below benchmark values that are associated with significant losses in fish 20 species richness. 21 Significant portions of the U.S. are acid sensitive, and current deposition levels exceed 22 those that would allow recovery of the most a cid sensitive lakes in the Adirondacks ( US EPA, 23Draft

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91 2008, Executive Summary). In addition, because of past loadings, areas of the Shenandoah are 1 sensitive to current deposition levels ( US EPA, 2008, Executive Summary). Parts of the West 2 are naturall y less se nsitive to acidification and subjected to lower deposition (particularl y SO x ) 3 levels relative to the eastern United States, and as such, less focus in the ISA is placed on the 4 adequacy

of the existing standards in these areas, with the exception of the mountainous areas of 5 the West, which experience episodic acidification due to deposition. 6 In describing the effects of acidification in the two case study areas the REA uses the 7 approach of describing benchmarks in terms of ANC values. Man y locations in sen sitive areas of 8 the U.S. have ANC levels below benchmark levels for ANC classified as severe, elevated, or 9 moderate concern ( US EPA, 20 11 , Figure 2

- 1). The average current ANC levels across 44 lakes 10 in the Adirondack case study area is 62.1 μeq/L (moderate concern). However, 44 percent of 11 lakes had deposition levels exceeding the critical load for an ANC of 50 μeq/L , and 28 percent of 12 lakes had deposition levels exceeding the critical load for an ANC of 20 μeq/L( US EPA, 200 9 , 13 section 4.2.4.2). This information indicates that almost half of the 44 lakes in the Adirondacks 14 case stud y area are at an elevated concern levels, and almost a third are at a severe concern level. 15 These levels are associated with greatl y diminished fish species diversit y, and losses i n the health 16 and reproductive capacity of remaining populations. Based on assessments of the relationship 17 between number of fish species and ANC level in both the Adirondacks and Shenandoah areas, 18 the number of fish species is decreased b y over half at an ANC level of 20 μeq/Lrelative to an 19 ANC level at 100 μeq/L ( US EPA, 200 9, Figure 4.2

- 1). When extrapolated to the full population 20 of lakes in the Adirondacks area using weights based on the EMAP probability survey ( US EPA, 21 200 9 , section 4.2.6.1 ) , 36 perc ent of lakes exceeded the critical load for an ANC of 50 μeq/Land 22 13 percent of lakes exceeded the critical load for an ANC of 20 μeq/L

. 23Draft

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92 Many streams in the Shenandoah case study area also have levels of deposition that are 1 associated with ANC levels cla ssified as severe, elevated, or moderate concern. The average 2 ANC under recent conditions for the 60 streams evaluated in the Shenandoah case study area is 3 57.9 μeq/L , indicating moderate concern. However, 85 percent of streams had recent deposition 4 exceed ing the critical load for an ANC of 50 μeq/L, and 72 percent exceeded the critical load for 5 an ANC of 20 μeq/L. As with the Adirondacks area, this information suggests that significant 6 numbers of sensitive streams in the Shenandoah area are at risk of adve rse impacts on fish 7 populations under recent conditions. Many other streams in the Shenandoah area are likely to 8 experience conditions of elevated to severe concern based on the prevalence in the area of 9 bedrock geology

associated with increased sensitivity to acidification suggesting that effects due 10 to stream acidification could be widespread in the Shenandoah area ( US EPA, 200 9 , section 11 4.2.6.2). 12 In addition to these chronic acidification effects, the I SA notes that “consideration of 13 episodic acidificat ion greatly increases the extent and degree of estimated effects for acidifying 14 deposition on surface waters.” ( US EPA, 2008, section 3.2.1.6 ) Some studies show that the 15 number of lakes that could be classified as acid

- impacted based on episodic acidification is 2 to 3 16 times the number of lakes classified as acid

- impacted based on chronic ANC. These episodic 17 acidification events can have long term effects on fish populations ( US EPA, 2008, section 18 3.2.1.6 ). Under recent conditions, episodic acidification has been observed in locations in the 19 eastern U.S. and in the mountainous western U.S. ( US EPA, 2008, section 3.2.1.6 ). 20 The I SA, REA and Policy Assessment all conclude that the current standards are not 21 adequate to protect against the adverse impacts of aquatic acidification on sensitive ecos ys tems. 22 In the IS A it is noted that significant portions of the U.S. are acid sensitive, and that current 23Draft

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93 deposition levels exceed those that would allow recovery of the most acid sensitive lakes in the 1 Adirondacks ( US EPA, 2008, Executive Summary). In addition, because of past loadings, areas 2 of the Shenandoah are sensitive to current deposition levels (US EPA, 2008, Executive 3 Summary). A recent survey , as reported in the ISA, found sensitive streams in many locatio ns in 4 the U.S., including the Appalachian Mountains, the Coastal Plain, and the Mountainous West 5 ( US EPA, 2008, section 4.2.2.3). In these sensitive areas, between 1 and 6 percent of stream 6 kilometers are chronically acidified. The REA further concludes th at both the Adirondack and 7 Shenandoah case stud y areas are currentl y receiving deposition from ambient oxides of nitrogen 8 and sulfur in excess of their ability to neutralize such inputs. In addition based on the current 9 emission scenarios, forecast modeli ng out to the year 2020 as well as 2050 indicates a large 10 number of streams in these areas will still be adversely impacted ( s ection IIB). Based on these 11 considerations, the Policy Assessment concludes that the current secondary NAAQS for oxides 12 of nitrog en and sulfur do not provide adequate protection of sensitive ecos ystems with regard to 13 aquatic acidification

. 14 ii. Terrestrial acidification 15 Based on the terrestrial acidification case studies (Kane Experimental Forest in 16 Pennsylvania and Hubbard Brook Expe rimental Forest (case study locations described in section 17 IIB) on sugar maple and red spruce habitat, the REA concludes that there is significant risk to 18 sensitive terrestrial ecos ystems from acidification at atmospheric concentrations of NO x and SO X 19 at o r below the current standards. The ecological indicator selected for terrestrial acidification 20 is the base cation to aluminum ratio (BC:Al), which has been linked to tree health and growth. 21 The results of the REA strongl y support a relationship between atmospheric deposition of oxides 22 of nitrogen and sulfur and BC:Al, and that BC:Al is a good indicator of terrestrial acidification. 23Draft

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94 At levels of deposition associated with oxides of nitrogen and sulfur concentrations at or below 1 the current standards, BC :Al levels are expected to be below benchmark values that are 2 associated with significant effects on tree health and growth. Such degradation of terrestrial 3
ecos ys tems could affect ecos ys tem services such as habitat provisioning, endangered species, 4 goods production (timber, sy rup, etc.) and many

others. 5 Many

locations in sensitive areas of the U.S. have BC:Al levels below benchmark levels 6 classified as providing low to intermediate levels of protection to tree health. At a BC:Al ratio of 7 1.2 (intermed iate level of protection), red spruce growth can be reduced b y 20 percent. At a 8 BC:Al ratio of 0.6 (low level of protection), sugar maple growth can be decreased b y 20 percent. 9 The REA did not evaluate broad sensitive regions. However, in the sugar map le case stud y area 10 (Kane Experimental Forest), recent deposition levels are associated with a BC:Al ratio below 11 1.2, indicating between intermediate and low level of protection, which would indicate the 12 potential for a greater than 20 percent reduction in growth. In the red spruce case stud y area 13 (Hubbard Brook Experimental Forest), recent deposition levels are associated with a BC:Al ratio 14 slightly

above 1.2, indicating slightl y better than an intermediate level of protection (US EPA, 15 200 9 , section 4.3.5.1 ). 16 Over the full range of sugar maple, 12 percent of evaluated forest plots exceeded the 17 critical loads for a BC:Al ratio of 1.2, and 3 percent exceeded the critical load for a BC:Al ratio 18 of 0.6. However, there was large variability across stat es. In New Jersey , 67 percent of plots 19 exceeded the critical load for a BC:Al ratio of 1.2, while in several states on the outskirts of the 20 range for sugar maple (e.g. Arkansas, Illinois) no plots exceeded the critical load for a BC:Al 21 ratio of 1.2. For red spruce, overall 5 percent of plots exceeded the critical load for a BC:Al ratio 22 of 1.2, and 3 percent exceeded the critical load for a BC:Al ratio of 0.6. In the major red spruce 23Draft

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95 producing states (Maine, New Hampshire, and Vermont), critical loads for a BC:Al ratio of 1.2 1 were exceeded in 0.5, 38, and 6 percent of plots. 2 The I SA, REA and Policy Assessment all conclude that the current standards are not 3 adequate to protect against the adverse impacts of terrestrial acidification on sensitive 4
ecos ys tems . The ISA reported one study that estimated 15 percent of U.S. forest ecos ys tems 5 exceeded the critical loads for acidit y for N and S deposition by >250 eq/ha/y ear under current 6 conditions ( US EPA, 2008, section 4.2.1.3 ). The main drawback, as stated in th e REA and 7 Policy

Assessment , with the understanding of terrestrial acidification lies in the availability of 8 data b y which we can predict critical loads and that the data are based on laboratory responses 9 rather than field measurements. Other stressors that are present in the field but that are not 10 present in the laboratory may confound this relationship. The REA does however, conclude that 11 the case stud y results suggest that nitrogen and sulfur acidifying deposition in the sugar maple 12 and red spruce forest areas studied were very close to, if not greater than, the critical loads for 13 those areas, and both ecosy stems are likel y to be sensitive to any future changes in the levels of 14 deposition. 15 iii. Terrestrial nutrient e nrichment 16 Nutrient enrichment effects are due to nitrogen loadings from both atmospheric and non

17 atmospheric sources. Evaluation of nutrient enrichment effects requires an understanding that 18 nutrient inputs are essential to ecosy stem health and that specific long term levels of nutrients in 19 a sy st em affect the t ypes of species that occur over long periods of time. Short term additions of 20 nutrients can affect species competition, and even small additions of nitrogen in areas that are 21 traditionally

nutrient poor can have significant impacts on produ ctivity as well as species 22 composition. Most ecosy stems in the United States are nitrogen

- limited, so regional decreases 23Draft

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96 in emissions and deposition of airborne nitrogen compounds could lead to some decrease in 1 growth of the vegetation that surrounds th e targeted aquatic s ys tem but as discussed below 2 evidence for this is mixed. Whether these changes in plant growth are seen as beneficial or 3 adverse will depend on the nature of the ecosy stem being assessed. 4 In formation on the effects of changes in nitro gen deposition on forestlands and other 5 terrestrial ecos ys tems is very limited. The multiplicity of factors affecting forests, including other 6 potential stressors such as ozone, and limiting factors such as moisture and other nutrients, 7 confound assessment s of marginal changes in any one stressor or nutrient in forest ecos ys tems. 8 The I SA notes that only a fraction of the deposited nitrogen is taken up by the forests, most of 9 the nitrogen is retained in the soils ( US EPA, 2008, section 3.3.2.1 ). In addition , the I SA 10 indicates that forest management practices can significantly affect the nitrogen cycling within a 11 forest ecos ys tem, and as such, the response of managed forests to NO x deposition will be 12 variable depending on the forest management practices emplo yed in a given forest ecos ys tem 13 ( US EPA, 2008, Annex C C.6.3 ). Increases in the availability of nitrogen in N

- limited forests 14 via atmospheric deposition could increase forest production over large non

- managed areas, but 15 the evidence is mixed, with some studies showing increased production and other showing little 16 effect on wood production ( US EPA, 2008, section 3.3.9) . Because leaching of nitrate can 17 promote cation losses, which in some cases create nutrient imbalances, slower growth and 18 lessened disease and freezing tolerances for forest trees, the net effect of increased N on forests 19 in the U.S. is uncertain ( US EPA, 2008, section 3.3.9). 20 The scientific literature has many examples of the deleterious effects caused by excessive 21 nitrogen loadings to terrestrial sy stems. Several studies have set benchmark values for levels of 22 N deposition at which scientifically adverse effects are known to occur. Large areas of the 23Draft

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97 country appear to be experiencing deposition above these benchmarks. The I SA presents that 1 studies have found that at 3.1 kg N/ha

- yr, the community of lichens begins to change from 2 acidoph yt ic to tolerant species; at 5.2 kg N/ha

- yr, the typical dominance b y acidoph yt ic species 3 no longer occurs; and at 10.2 kg N/ha

- yr, acidophytic lichens are totall y lost from the 4 community . Additional studies in the Colorado Front Range of the Rocky Mountain National 5 Park support these findings. These three values (3.1, 5.2, and 10.2 kg/ha

- yr ) are one set of 6 ecologicall y meaningful benchmarks for the mixed conifer forest (MCF) of the pacific coast 7 regions. Nearl y all of the known sensitive communities receive total nitrogen deposition levels 8 above the 3.1 N kg/ha

- yr ecological benchmark according to the12 km, 2002 CMAQ/NADP 9 data, with the exception of the easternmost Sierra Nevadas. MCFs in the southern portion of the 10 Sierra Nevada forests and nearl y all MCF communities in the San Bernardino forest s receive 11 total nitrogen deposition levels above the 5.2 N kg/ha

- yrecological benchmark. 12 Coastal Sage Scrub communities (CSS) are also known to be sensitive to community 13 shifts caused b y excess nitrogen loadings. Studies have investigated the amount of nitrogen 14 utilized by

healthy and degraded CSS sy stems. In healthy stands, the authors estimated that 3.3 15 kg N/ha

- yr was used for CSS plant growth. It is assumed that 3.3 kg N/ha

- yr is near the point 16 where nitrogen is no longer limiting in the CSS community. Therefore, this amount can be 17 considered an ecological benchmark for the CSS community . The majority of the known CSS 18 range is currentl y receiving deposition in excess of this benchmark. Thus, the REA concludes 19 that recent conditions where oxides of ni trogen ambient concentrations are at or below the 20 current oxides of nitrogen secondary standards are not adequate to protect against anticipated 21 adverse impacts from N nutrient enrichment in sensitive ecosy stems. 22Draft

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98 iv. Aquatic nutrient e nrichment 1 The REA aquatic nutrient enrichment case studies focused on coastal estuaries and 2 revealed that while current ambient loadings of atmospheric oxides of nitrogen are contributing 3 to the overall depositional loading of coastal estuaries, other non

- atmospheric sources ar e 4 contributing in far greater amounts in total, although atmospheric contributions are as large as 5 some other individual source t ypes. The ability of current data and models to characterize the 6 incremental adverse impacts of nitrogen deposition is limited , both by the available ecological 7 indicators, and b y the inability to attribute specific effects to atmospheric sources of nitrogen. 8 The REA case studies used as the ecological indicator for aquatic nutrient enrichment an index 9 of eutrophication known as the Assessment of Estuarine Trophic Status Eutrophication Index 10 (ASSETS EI ). This index is a six level index characterizing overall eutrophication risk in a 11 waterbod y. This indictor is not sensitive to relatively large changes in nitrogen deposition. In 12 addition, this ty

pe of indicator does not reflect the impact of nitrogen deposition in conjunction 13 with other sources of nitrogen. 14 Based on the above considerations, the REA concludes that the ASSETS EI is not an 15 appropriate ecological indicator for e stuarine aquatic eutrophication and that additional analy sis 16 is required to develop an appropriate indicator for determining the appropriate levels of 17 protection from N nutrient enrichment effects in estuaries related to deposition of oxides of 18 nitrogen

. As a result, EPA is unable to make a determination as to the adequacy of the existing 19 secondary oxides of nitrogen standard in protecting public welfare from N nutrient enrichment 20 effects in estuarine aquatic ecos ys tems. 21 Additionally , nitrogen deposition can alter species composition and cause eutrophication 22 in freshwater s ys tems. In the Rocky Mountains, for example, deposition loads of 1.5 to 2 kg/ha

- 23Draft

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99 yrwhich are well within current ambient levels are known to cause changes in species 1 composition in diat om communities indicating impaired water qualit y ( US EPA, 2008, section 2 3.3.5.3 ). It then seems apparent then that the existing secondary standard for oxides of nitrogen 3 does not protect such ecosy stems and their resulting services from impairment. 4
v. Other effects 5 An important consideration in looking at the effects of deposition of oxides of sulfur in 6 aquatic ecosy stems is the potential for production of methylmercury , a highly neurotoxic 7 contaminant . The production of meaningful amounts of methylmercury (MeHg) requires the 8 presence of SO 4 2
- and mercury , and where mercury is present, increased availability of SO 4 2
- 9 results in increased production of MeHg. There is increasing evidence on the relationship 10 between sulfur deposition and increased meth yl ation o f mercury in aquatic environments; this 11 effect occurs only

where other factors are present at levels within a range to allow methylation. 12 The production of meth ylmercury requires the presence of sulfate and mercury , but the amount 13 of methylmercury

produced varies with ox ygen content, temperature, pH, and supply of labile 14 organic carbon ( US EPA, 2008, section 3.4). In watersheds where changes in sulfate deposition 15 did not produce an effect, one or several of those interacting factors were not in the range 16
re quired for meaningful methylation to occur ( US EPA, 2008, section 3.4). Watersheds with 17 conditions known to be conducive to mercury methylation can be found in the northeastern 18 United States and southeastern Canada ( US EPA, 200 9 , section 6). 19 With respect to sulfur deposition and mercury methylation, the final ISA determined: The 20 evidence is sufficient to infer a causal relationship between sulfur deposition and increased 21 mercury

methylation in wetlands and aquatic environments. However, EPA did not conduc t a 22 quantitative assessment of the risks associated with increased mercury methylation under current 23Draft

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100 conditions. As such, EPA is unable to make a determination as to the adequacy of the existing 1 SO 2 secondary standards in protecting against welfare effects associated with increased mercury 2 methylation. 3 vi. Summary of Adequacy Considerations 4 In summary , the Policy Assessment concludes that currently available scientific evidence 5 and assessments clearl y call into question the adequacy of the current standard s with regard to 6 deposition

- related effects on sensitive aquatic and terrestrial ecos ys tems, including acidification 7 and nutrient enrichment. Further, the Policy Assessment recognizes that the elements of the 8 current standards

-- indicator, averaging time , level and form

– are not ecologically relevant, and 9 are thus not appropriate for standards designed to provide such protection. Thus, the Policy 10 Assessment concludes that consideration should be given to establishing a new ecologically 11 relevant multi

- po llutant, multimedia standard to provide appropriate protection from deposition

12 related ecological effects of oxides of nitrogen and sulfur on sensitive ecosystems, with a focus 13 on protecting against adverse effects associated with acidifying deposition in sensitive aquatic 14
ecos ys tems. 15 3. CASAC Views 16 In a letter to the Administrator (Russell and Samet 2011 a), the CASAC Oxides of 17 Nitrogen and Oxides of Sulfur Panel, with full endorsement of the chartered CASAC, 18 unanimously

concluded that : 19 EPA staff has demonstrated through the Integrated Science Assessment (I SA), 20 Risk and Exposure Characterization (REA) and the draft PA that ambient NOx 21 and SOx can have, and are having, adverse environmental impacts. The Panel 22 views that the current NOx and SOx secondary standards should be retained to 23 protect against direct adverse impacts to vegetation from exposure to gas phase 24 exposures of these two families of air pollutants. Further, the ISA, REA and draft 25 PA demonstrate that adverse impacts to aquatic ecosy stems are also occurring due 26 to deposition of NOx and SOx. Those impacts include acidification and 27Draft

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101 undesirable levels of nutrient enrichment in some aquatic ecosy stems. The levels 1 of the current NOx and SOx secondary NAAQS are not sufficient, nor the forms 2 of tho se standards appropriate, to protect against adverse depositional effects; thus 3 a revised NAAQS is warranted. 4 5 Inaddition, with regard to the joint consideration of both oxides of nitrogen and oxides 6 of sulfur as well as the consideration of deposition related effects, CASAC concluded that the 7 Policy

Assessment had developed a credible methodology for considering such effects. The 8 Panel stated that “the Policy Assessment develops a framework for a multi

- pollutant, multimedia 9 standard that is ecologicall y relevant and reflects the combined impacts of these two pollutants as 10 they

deposit to sensitive aquatic ecosy stems.” 11 4. Administrator’s Proposed Conclusions Concerning Adequacy of Current Standard 12 Based on the above considerations and taking into account C ASAC advice, the 13 Administrator recognizes that the purpose of the secondary standard is to protect against 14 “adverse” effects resulting from exposure to oxides of nitrogen and sulfur, discussed above in 15 section IIA. The Administrator also recognizes the nee d for conclusions both to the adequacy of 16 the current standards for both direct and deposition related effects as well as conclusions as to the 17 appropriateness and ecological relevance of the current standards. 18 In considering what constitutes an ecologi cal effect that is also adverse to the public 19 welfare, the Administrator took into account the ISA conclusions regarding the nature and 20 strength of the effects evidence, the risk and exposure assessment results, the degree to which the 21 associated uncertainties should be considered in interpreting the results, the conclusions 22 presented in the Policy Assessment , and the views of CASAC and members of the public. On 23 these bases, the Administrator concludes that the current secondary standard s are adequate to 24 p rotect against direct ph ytotoxic effects on vegetation. Thus the Administrator proposes to retain 25Draft

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102 the current secondary

standards for oxides of nitrogen at 0.053 ppm as an annual average and the 1 current secondary

standards for oxides of sulfur at 0.5 ppm as a 3 hour average. 2 With regard to deposition

- related effects, the Administrator has first to consider the 3 appropriateness of the structure of the current standards to address ecological effects of concern. 4 Based on the evidence as well as considerin g the advice given b y CASAC on this matter, the 5 Administrator conclud es that the elements of the current standards are not ecologically relevant 6 and thus are not appropriate to provide protection of ecos ys tems
. On the subject of adequacy of 7 protection with regard to deposition

- related effects, the Administrator considered the full nature 8 of ecological effects related to the deposition of ambient oxides of nitrogen and sulfur into 9 sensitive ecos ys tems across the country . Her conclusions are based on the evi dence presented in 10 the I SA with regard to acidification and nutrient enrichment effects, the findings of the REA with 11 regard to scope and severity of the current and likely

future effects of deposition, the sy nthesis of 12 both the scientific evidence and ris k and exposure results in the Policy Assessment as to the 13 adequacy

of the current standards, and the advice of both CASAC and the public. After such 14 consideration, the Administrator concludes that current levels of oxides of nitrogen and sulfur are 15 suffi cient to cause acidification of both aquatic and terrestrial ecos ys tems, nutrient enrichment of 16 terrestrial ecos ys tems and contribute to nutrient enrichment effects in estuaries that could be 17 considered adverse, and the current secondary standards do not p rovide adequate protection from 18 such effects. 19 Having reached these conclusions, the Administrator determines that it is appropriate to 20 consider alternative standards that are ecologically relevant. These considerations support the 21 conclusion that the c urrent secondary standard s is neither appropriate nor adequate to protect 22Draft

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103 against deposition related effects

. The Administrator’s consideration of such alternative 1 standards is discussed below in Section III. 2 III. Rationale for Proposed Decision on Alternative Secondary Standards for Aquatic 3 Acidification 4 Having reached the conclusion that the current NO 2 and SO 2 secondary standards are not 5 adequate to provide appropriate protection against deposition

- related effects associated with 6 oxides of nitrogen and sulfur, the Administrator then considered what new standard might be 7 appropriate, at this time, to address suc h effects on public welfare. The Administrator recognizes 8 that the inherentl y complex and variable linkages between ambient concentrations of nitrogen 9 and sulfur oxides, the related deposited forms of nitrogen and sulfur, and the ecological 10 responses that are associated with public welfare effects call for consideration of an ecologically 11 relevant design of a standard that reflects these linkages. The Administrator also recognizes that 12 characterization of such complex and variable linkages will necessaril y require consideration of 13 information and anal ys es that have important limitations and uncertainties. 14 Despite its complexity , an ecologicall y relevant multi

- pollutant standard to address 15 deposition

- related effects could still appropriately be defined in terms of the same basic elements 16 that are used to define any NAAQS

– indicator, form, averaging time, and level. The form 17 would incorporate additional structural elements that reflect relevant multi

- pollutant and 18 multimedia attributes. These structural e lements include the use of an ecological indicator, tied 19 to the ecological effect we are focused on, and other elements that account for ecologically 20 relevant factors other than ambient air concentrations. All of these elements would be needed to 21 enable a linkage from ambient air indicators to the ecological indicator to define an ecologicall y 22 relevant standard. As a result, such a standard would necessaril y be more complex than the 23Draft

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104 NAAQS that have been set historically to address effects associated with ambient concentrations 1 of a single pollutant. 2 More specificall y, the Administrator considered an ecologicall y relevant standard to 3 address effects associated with acidifying deposition related to ambient concentrations of oxides 4 of nitrogen and sulfur in sensitive aquatic ecosy stems. This focus is consistent with the 5 information presented in the I SA, REA, and Policy Assessment which highlighted the 6 sufficiency

of the quantity

and quality

of the available evidence and assessments associated with 7 aquatic ac idification relative to the information and assessments available for other deposition

8 related effects, including terrestrial acidification and aquatic and terrestrial nutrient enrichment. 9 Based on its review of these documents, CASAC agreed that aquatic acidification should be the 10 focus for developing a new multi

- pollutant standard in this review. In reaching conclusions 11 about an air qualit y standard designed to address deposition

- related aquatic acidification effects, 12 the Administrator also recognizes that such a standard may also provide some degree of 13 protection against other deposition

- related effects. 14 As discussed in chapter 7 of the Policy Assessment, the development of a new standard to 15 address deposition

- related aquatic acidification effects reco gnizes the need for consideration of a 16 nationally

applicable standard for protection against adverse effects of aquatic acidification to 17 public welfare, while recognizing the complex and heterogeneous interactions between ambient 18 air concentrations of nitr ogen and sulfur oxides, the related deposition of nitrogen and sulfur, and 19 associated ecological responses. The development of such a standard also needs to take into 20 account the limitations and uncertainties in the available information and analyses upon which 21 characterization of such interactions are based. The approach used in the Policy Assessment also 22 recognizes that while such a standard would be national in scope and coverage, the effects to 23Draft

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105 public welfare from aquatic acidification will not occur to the same extent in all locations in the 1 U.S., given the inherent variability of the responses of aquatic s ys tems to the effects of acidifying 2 deposition. 3 As discussed above in section II, man y locations in the U.S. are naturall y protected 4 against acid deposition due to underl yi ng geological conditions. L ikewise, some locations in the 5 U.S., including lands managed for commercial agriculture and forestry , are not likely to be 6 negativel y impacted b y current levels of nitrogen and sulfur deposition. As a result, while a new 7 ecologicall y relevant secondary standard would apply everywhere, it would be structured to 8 account for differences in the sensitivity of ecos ystems across the country. This would allow for 9 appropriate protection of sensitive aquatic ec osy stems, which are relativel y pristine and wild and 10 generall y in rural areas, and the services provided by such sensitive ecos ys tems, without 11 requiring more protection than is needed elsewhere. 12 As discussed below, the standard developed in the Policy A ssessment would employ (1) 13 NOy

and SOx as the atmospheric ambient air indicators; (2) a form that takes into account 14 variable factors, such as atmospheric and ecos ys tem conditions that modify the amounts of 15 deposited nitrogen and sulfur; the distinction be tween oxidized and reduced forms of nitrogen; 16 effects of deposited nitrogen and sulfur on aquatic ecos ys tems in terms of the ecological 17 indicator ANC; and the representativeness of water bodies within a defined spatial area; (3) a 18 multi

- year averaging time , and (4) a standard level defined in terms of a single, national target 19 ANC value that, in the context of the above form, identifies the levels of concentrations of NOy 20 and SOx in the ambient air that would meet the standard. The form of such a standard has been 21 defined b y an index, termed an aquatic acidification index (AAI), which reflects the relationship 22Draft

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106 between ambient concentrations of NOy and SOx and aquatic acidification effects that result 1 from nitrogen and sulfur deposition related to these ambi ent concentrations. 2 In presenting the considerations associated with such an air quality standard to address 3 deposition

- related aquatic acidification effects, the following sections focus on each element of 4 the standard, including indicator (section III

. A ), form (section III

. B), averaging time (section 5 III
. C), and level (section III

. D). Alternative combinations of levels and forms are discussed in 6 section III.E. Considerations related to important uncertainties inherent in such an approach are 7 discussed in section III

. F. Advice from CASAC on such a new standard is presented in section 8 III.G. The Administrator’s proposed decisions on such a new standard are presented in section 9 III.H. 10 A. Ambient Air Indicators 11 In considering alternative ambient air ind icators, the Policy Assessment primarily

focuses 12 on the important attribute of association. Association in a broad sense refers to how well an 13 ambient air indicator relates to the ecological effects of interest by virtue of both the framework 14 that links the ambient indicator and effects and the empirical evidence that quantifies the 15 linkages. The Policy Assessment also considers how measurable or quantifiable an indicator is 16 to enable its use as an effective indicator of relevant ambient air concentration s. 17 As discussed above in section II.C, the Policy Assessment concludes that indicators other 18 than NO 2 and SO 2 should be considered as the appropriate indicators of oxides of nitrogen and 19 sulfur in the ambient air for protection against the acidification e ffects associated with deposition 20 of the associated nitrogen and sulfur. This conclusion is based on the recognition that all forms 21 of nitrogen and sulfur in the ambient air contribute to deposition and resulting acidification, and 22 as such, NO 2 and SO 2 ar e incomplete indicators. In principle, the ambient indicators should 23Draft

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107 represent the species that are associated with oxides of nitrogen and sulfur in the ambient air and 1 can contribute acidifying deposition. This includes both the species of oxides of nitrogen and 2 sulfur that are directl y emitted as well as species transformed in the atmosphere from oxides of 3 nitrogen and sulfur that retain the nitrogen and sulfur atoms from directly emitted oxides of 4 nitrogen and sulfur. All of these compounds are assoc iated with oxides of nitrogen and sulfur in 5 the ambient air and can contribute to acidifying deposition. 6 The Policy

Assessment focuses in particular on the various compounds with nitrogen or 7 sulfur atoms that are associated with oxides of nitrogen and s ulfur, because the acidifying 8 potential is specific to nitrogen and sulfur, and not other atoms (e.g., H, C, O) whether derived 9 from the original source of oxides of nitrogen and sulfur emissions or from atmospheric 10 transformations. For example, the acidi fying potential of each molecule of NO 2 , NO, HNO 3 or 11 PAN is identical, as is the potential for each molecule of SO 2 or ion of particulate sulfate, p - SO 4 . 12 Each atom of sulfur affords twice the acidifying potential of each atom of nitrogen. 13 1. Oxides of Su lfur 14 As discussed in the Policy Assessment ( US EPA, 2011, section 7.1.1), oxides of sulfur 15 include the gases sulfur monoxide (SO), sulfur dioxide (SO 2 ), sulfur trioxide (SO 3 ), disulfur 16 monoxide (S 2 O), and particulate

- phase sulfur compounds that result from gas

- phase sulfur 17 oxides interacting with particles. However, the sum of SO 2 and SO 4 does represent virtuall y the 18 entire ambient air mass of sulfur that contributes to acidification. In addition to accounting for 19 virtually

all the potential for acidi fication from oxidized sulfur in the ambient air, there are 20 reliable methods to monitor the concentrations of SO 2 and particulate SO 4 . In addition, much of 21 the data used to develop the technical basis for the standard developed in the Policy Assessment 22Draft

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108 is based on monitoring or modeling of these species. 4 The Policy Assessment concludes that the 1 sum of SO 2 and SO 4 , referred to as SOx, are appropriate ambient air indicators of oxides of 2 sulfur because they

represent virtually

all of the acidification potential of ambient air oxides of 3 sulfur and there are reliable methods suitable for measuring SO 2 and SO 4 . 4 2. Oxides of Nitrogen 5 As discussed in the Policy Assessment ( US EPA, 2011, section 7.1.2), total reactive 6 oxidized nitrogen, NOy , as defined in chapter 2 of the Policy Assessment, incorporates basicall y 7 all of the oxidized nitrogen species that have acidifying potential and as such, NOy should be 8 considered as an appropriate indicator for oxides of nitrogen. NOy is an aggregate measure of s 9 NO and NO 2 an d all of the reactive oxidized products of NO and NO 2 . That is, NOy is a group 10 of nitrogen compounds in which all of the compounds are either an oxide of nitrogen or 11 compounds in which the nitrogen atoms came from oxides of nitrogen. NOy is especiall y 12
re levant as an ambient indicator for acidification in that it both relates to the oxides of nitrogen in 13 the ambient air and also represents the acidification potential of all oxidized nitrogen species in 14 the ambient air, whether an oxide of nitrogen or deriv ed from oxides of nitrogen. 15 There are currently available reliable methods of measuring aggregate NOy . The term 16 “aggregate” measure means that the NO Y , as measured, is not based on measuring each 17 individual species of NOy and calculating an NOy value b y summing the individual species. 18 Rather, as described in chapter 2 of the Policy Assessment, current measurement techniques 19 process all of the individual NOy species to produce a single aggregate measure of all of the 20 nitrogen atoms associated with any N Oy species. Consequently , the NOy measurement 21

4 As discussed in chapter 2 of the P olicy A ssessment , SO 2 and particulate SO 4 are routinely measured in ambient air mo nitoring networks, although only CASTNET filter packs do not intentionally exclude particle size fractions. The CMAQ treatment of SOx is the simple addition of both species, which are treated explicitly in the model formulation. All particle size fractions are included in the CMAQ SOx estimates.Draft

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109 effectivel y provides the sum of all individual species, but the identity of the individual species is 1 lost. As discussed above, the accounting for the individual nitrogen atoms is an accounting of 2 the ambient air acidification potential of oxides of nitrogen and their transformation products and 3 therefore the most relevant ambient indicator for aquatic acidification effects associated with 4 oxides of nitrogen. 5 This loss of the information on individual spec ies motivated consideration of alternative 6 or more narrowl y defined indicators for oxides of nitrogen in the Policy Assessment. 7 Consideration of a subset of NOy species was based on the following reasoning. First, the actual 8 dry deposition of nitrogen is determined on an individual species basis b y multiplying the species 9 concentration times a species

- specific deposition velocity and then summed to develop an 10 estimate of total dry deposition. Consequently , the use of individual ambient species has the 11 po tential to be more consistent with the underlying science of deposition and, therefore, has the 12 potential to allow for a more rigorous evaluation of dry deposition with specialized field studies. 13 In addition, there has been a suggestion of focusing onl y o n the most quickly depositing NOy 14 species, such as HNO 3 , as contributions from other NOy species such as NO 2 may be negligible. 15 These alternative indicators are discussed below. 16 The Policy

Assessment considers the relative merits of using each individual NOy 17 species as part of a group of indicators. In so doing, it was first noted that dry deposition of NOy 18 is treated as the sum of the deposition of each individual species in advanced process

- based air 19 quality

models like CMAQ, as described in chapter 2 of the Policy Assessment. Conceptually 20 one could extend this process

- based approach b y using all NO Y species individually as separate 21 indicators for oxides of nitrogen and requiring, for example, measurements of each of the 22 species, including the dominant species of HNO 3 , particulate nitrate (p

- NO 3 ) , true NO 2 , NO, and 23Draft

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110 PAN. The potential attraction of using individual species would be the reliance on actual 1 deposition velocities. This could have more ph ys ical meaning in comparison to a constructed 2 model of aggregate deposition of NOy , which is difficult to evaluate with observations because 3 of the assimilation of many species with disparate deposition behavior. The Policy Assessment 4 notes that the major drawback of using individual NOy species as the indicators is the lack of 5 reliable measurement techniques, especiall y for PAN and NO 2 in rural locations, which renders 6 the use of virtuall y an y i ndividual NOy species, except for NO and perhaps p - NO 3 , as 7 functionally

inadequate from a measurement perspective

. 8 The Policy

Assessment next considered the relative merits of using a subset of NOy 9 species as the indicators for oxides of nitrogen, as was discussed above for oxides of sulfur. To 10 the extent that certain species provide relativel y minor contributions to total NOy deposition, it 11 may

be appropriate to consider excluding them as part of the indicator. As discussed in chapter 2 12 of the Policy

Assessment, each nitrogen species within the array of NOy species has species

13 specific dry

deposition velocities. For example, the deposition velocity of HNO 3 is much 14 greater than the velocit y for nitrogen dioxide and, consequentl y, for a similar ambient air 15 concentration, HNO 3 contributes more deposition of acidifying nitrogen relative to nitrogen 16 dioxide. In trans itioning from source

- oriented urban locations to rural environments, the ratio of 17 the concentrations of HNO 3 and PAN to NO 2 increases. 18 Based on the reasoning that a larger fraction of the deposited NO Y is accounted for by 19 total nitrate (the sum of HNO 3 an d p - NO 3 ), a surrogate for the more rapidly depositing fraction 20 of NOy , combined with the availability of reliable total nitrate measurements through the Clean 21 Air Status and Trends Network ( CASTNET ) , the Policy Assessment considered using total 22 nitrate as the indicator for oxides of nitrogen ( US EPA, 2011, appendix E ). Nitrate would be 23Draft

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111 expected to correlate well with total reactive oxidized nitrogen deposition relative to NO Y ( US 1 EPA, 2011, chapter 2 ) , gi ven the inherent noise associated with variable contributions of low 2 deposition velocity

species (e.g., NO 2 ) that may have relatively

high ambient concentrations. 3 However, modeling simulations suggest that NOy may be a more robust indicator, relative to 4 HNO 3 , in terms of relating absolute changes in ambient air concentrations to changes in nitrogen 5 deposition driven by changes in ambient concentrations of oxides of nitrogen ( US EPA, 6 2011, Figure 2
- 32 ). 7 Based on the above considerations, the Policy Assessment c oncludes that total reactive 8 oxidized nitrogen, NOy , should be considered as the appropriate ambient indicator for oxides of 9 nitrogen based on its direct relationship to oxides of nitrogen in the ambient air and its direct 10 relationship to deposition associ ated with aquatic acidification. Because NOy represents all of 11 the potentially

acidifying oxidized nitrogen species in the ambient air, it is appropriatel y 12 associated with the deposition of potentially acidifyingcompounds associated with oxides of 13 nitrog en in the ambient air. In addition, there are reliable methods available to measure NOy

. 14 Measurement of each individual species of NOy , or the measurement of only a subset of species 15 of NOy , is less appropriate because there are not reliable measurements methods available to 16 measure all of the individual species of NOy and a subset of species would fail to account for 17 significant portions of the oxidized reactive nitrogen that relates to acidification. 5 18 B. Form 19 Based on the evidence of the aquatic acidif ication effects caused by

the deposition of 20 NOy

and SOx, the Policy

Assessment ( US EPA, 2011, section 7.2) presents the development of a 21 new form that is ecologically relevant for addressing such effects. The conceptual design for the 22

5 The P olicy A ssessment also note s that NO Y is a useful measurement for model evaluation purposes , which is especially imp ortant, recognizing the unique role that CMAQ play s in the development of this standard, as described below in section III.B
.Draft

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112 form of such a standard includes three main components: an ecological indicator, deposition 1 metrics that relate to the ecological indicator, and a function that relates ambient air indicators to 2 deposition metrics. Collectivel y, these three components link the ecological ind icator to ambient 3 air indicators, as illustrated above in Fig II
- 1. 4 The simplified flow diagram in Figure II
- 1 compresses the various atmospheric, 5 biological, and geochemical processes associated with acidifying deposition to aquatic 6
ecos ys tems into a simplified conceptual picture. The ecological indicator (left box) is related to 7 atmospheric deposition through biogeochemical ecosy stem models (middle box), which 8 associate a target deposition load to a target ecological indicator. Once a target depositio n is 9 established, associated allowable air concentrations are determined (right box) through the 10 relationships between concentration and deposition that are embodied in air quality models such 11 as CMAQ. The following discussion describes the development and rationale for each of these 12 components, as well as the integration of these components into the full expression of the form 13 of the standard using the concept of a national AAI that represents a target ANC level as a 14 function of ambient air concentration s. Spatial aggregation issues associated with defining each 15 of the terms of this index are also addressed below. 16 The AAI is designed to be an ecologically relevant form of the standard that determines 17 the levels of NOy

and SOx in the ambient air that wo uld achieve a target ANC limit for the U.S. 18 The intent of the AAI is to weight atmospheric concentrations of oxides of nitrogen and sulfur b y 19 their propensity

to contribute to acidification through deposition, given the fundamental 20 acidifying potential of each pollutant, and to take into account the ecological factors that govern 21 acid sensitivity

in different ecosy stems. The index also accounts for the contribution of reduced 22 nitrogen to acidification. Thus, the AAI encompasses those attributes of specif ic relevance to 23Draft

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113 protecting ecos ys tems from the acidifying potential of ambient air concentrations of NOy and 1 SOx. 2 1. Ecological Indicator 3 In considering alternative ecological indicators, the Policy Assessment again primarily 4 focuses on the attribute of association. In the case of an ecological indicator for aquatic 5 acidification, association refers to the relationship between the indicator and adverse effects as 6 discussed in section II. Because of the conceptual structure of the form of this standard (Figure 7 III
- 1), this particular ecological indicator must also link up in a meaningful and quantifiable 8 manner with acidifying atmospheric de position. In effect, the ecological indicator for aquatic 9 acidification is the bridge between biological impairment and deposition of NOy and SOx. 10 This section presents the rationale in the Policy Assessment for selecting acid 11 neutralizing capacit y ( ANC) as the appropriate ecological indicator for consideration. 12 Recognizing that ANC is not itself the causative or toxic agent for adverse aquatic acidification 13 effects, the rationale for using ANC as the relevant ecological indicator is based on the 14 fol lowing: 15 (1) ANC is directly associated with the causative agents, pH and dissolved aluminum, both 16 through empirical evidence and mechanistic relationships; 17 (2) Empirical evidence shows very clear and strong relationships between adverse effects 18 and ANC; 19 (3) ANC is a m ore reliable indicator from a modeling perspective, allowing use of a body of 20 studies and technical analy ses related to ANC and acidification to inform the 21 development of the standard; and 22Draft

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114 (4) ANC literally

embodies the concept of acidification as posed by the basic principles of 1 acid base chemistry and the measurement method used to estimate ANC and, therefore, 2 serves as a direct index to protect against acidification. 3 4 Ecological indicators of acidification in aquatic ecosy stems can be chemical or biologic al 5 components of the ecosystem that are altered b y the acidifying effects of nitrogen and sulfur 6 deposition. A desirable ecological indicator for aquatic acidification is one that is measurable or 7 estimable, linked causally to deposition of nitrogen and su lfur, and linked causally , either 8 directly

or indirectl y to ecological effects known or anticipated to adversely affect public 9 welfare. 10 As summarized in chapter 2 of the Policy Assessment, atmospheric deposition of NOy 11 and SOx causes aquatic acidification through the input of strong acid anions (e.g., NO 3 - and 12 SO 4 2
- ) that ultimately shifts the water chemistry equilibrium toward increased hydrogen ion 13 levels (or decreased pH). The anions are deposited either directly to the aquatic ecos ys tem or 14 indirectly via transformation through soil nitrification processes and subsequent drainage from 15 terrestrial ecos ys tems. In other words, when these anions are mobilized in the terrestrial soil, 16 they

can leach into adjacent water bodies. Aquatic acidification is indi cated by changes in the 17 surface water chemistry of ecos ys tems. In turn, the alteration of surface water chemistry has been 18 linked to negative effects on the biotic integrit y of freshwater ecos ys tems. There is a suite of 19 chemical indicators that could be us ed to assess the effects of acidifying deposition on lake or 20 stream acid

- base chemistry . These indicators include acid neutralizing capacity (ANC); 21 alkalinity

(ALK); base neutralizing capacit y, commonly referred to as acidity

(ACY); surface 22Draft

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115 water pH; conce ntrations of trivalent aluminum, Al +3 ; and concentrations of major anions (SO 4 2
- , 1 NO 3 - ), cations (Ca2+ , Mg +2 , K+ ), or sums of cations or anions. 2 ANC and ALK are very similar quantities and are used interchangeabl y in the literature 3 and for some of the ana ly ses presented in this document. Both ANC and ALK are defined as the 4 amount of strong acid required to reach a specified equivalence point. For acid

- base solutions, 5 an equivalence point can be thought of as the point to which the addition of strong aci ds (i.e., 6 titration) is no longer neutralized by the solution. This explains the term acid neutralizing 7 capacity , or ANC, as ANC relates directly to the capacit y of a s ys tem to neutralize acids. The 8 differences between ANC and ALK are based on operation al definitions and subject to various 9 interpretations. ANC is preferred over ALK as the body of scientific evidence has focused on 10 ANC and effects relationships. ALK is more widely associated with more general 11 characterizations of water quality such as the relative hardness of water associated with 12 carbonates. 13 Indictors such as the concentrations of specific anions, cations, or their groupings, while 14 relevant to acidification processes, are not robust acidification indicators as it is the relative 15 balance of cations and anions that is more directly associated with acidification. That balance is 16 captured b y ANC and ALK. Acidity, ACY, is the corollary of ANC from the perspective of 17 how much strong base it takes to reach an equivalence point. Becau se ACY is not used in most 18
ecos ys tem assessments, the body of information relating ACY to effects is too limited to serve as 19 a basis for an appropriate ecological indicator. Aluminum and other metals are causative toxic 20 agents that directl y impair biologi cal functions. However, aluminum, or metals in general, have 21 high variability

in concentrations that can be linked to effects, often at extremely low levels 22 which in some cases approach detectability limits, exhibit rapid transient responses, and are ofte n 23Draft

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116 confounded b y the presence of other toxic metals. These concerns limit the use of metals as 1 reliable and measurable ecological indicators. Hydrogen ion (H + ) concentrations, using their 2 negative logarithmic values, or pH, are well correlated with adverse effects, as discussed above 3 in section II.A, and determine the solubility of metals such as aluminum. However, pH is not a 4 preferred acidification indicator due to its highl y transient nature and other concerns, as 5 discussed below. 6 Having reasoned th at ANC is a preferred indicator to ALK, ACY, individual metals or 7 groupings of ions, the Policy Assessment considers the relative merits of ANC compared to pH, 8 which is a well recognized indicator of acidity and a more direct causative agent with regard to 9 adverse effects. First, the linkage between ANC and pH is considered in recognition of the 10 causative association between pH and effects. 11 ANC is not the direct causative toxic agent impacting aquatic species diversity . The 12 scientific literature general ly emphasizes the links between pH and adverse effects as described 13 above in section II.A. It is important, therefore, to consider the extent to which ANC and pH are 14 well related from a mechanistic perspective as well as through empirical evidence. ANC and 15 pH are co
- dependent on each other based on the requirement that all solutions are electrically 16 neutral, meaning that an y solution must satisfy the condition that all negatively charged species 17 must be balanced b y all positively charged species. ANC is defined as the difference between 18 strong anions and cations ( US EPA, 2011 , equation 7

- 13). 19 While the chemistry can be complex, the co

- dependency between ANC and pH is 20 explained by

recognizing that positively charged hydrogen, H + , is incorporated in the charge 21 balance relationships related to the overall solution chemistry which also defines ANC. The 22 positive, directional co

- dependency (i.e., ANC and pH increase together) is further explained in 23Draft

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117 concept as ANC reflects how much strong acid (i.e., how mu ch h ydrogen ion) it takes to titrate to 1 an equivalence point. Strong observed correlations between pH and ANC as described in the 2 Policy

Assessment support these mechanistic relationships. 3 As discussed above in section II.A, there are well established examples of ANC 4 correlating strongl y with a variet y of ecological effects which are summarized in the Policy 5 Assessment ( US EPA, 2011, Table 3

- 1). Because pH and ANC are well correlated and linearl y 6 dependent over the pH ranges (4.5

- 6) where adverse ecol ogical effects are observed, evidence of 7 clear associations exist between ANC and adverse ecological effects as described in the Policy 8 Assessment. In large measure, this dependence between pH and ANC and the relationship of 9 both pH and ANC to effects, sp eak directly to the appropriateness of ANC with respect to its use 10 as an ecological indicator. 11 Thus, there is a clear association between ANC and ecological effects, although there is a 12 more direct causal relationship between pH and effects. Nonetheless, ANC is preferred as an 13 ecological indicator based on its superior ability to provide a linkage with deposition in a 14 meaningful and quantifiable manner, a role that is served far more effectively by ANC than b y 15 pH. While both ANC and pH are clearly associ ated with the effects of concern, ANC is superior 16 in linking these effects to deposition. 17 The Policy

Assessment notes that the basis for this conclusion is that acidifying 18 atmospheric deposition of nitrogen and sulfur is a direct input of potential acid ity (ACY), or, in 19 terms of ANC, such deposition is relevant to the major anions that reduce the capacity of a water 20 body

to neutralize acidity. Consequently , there is a well defined linear relationship between 21 potential acidifying deposition and ANC. Thi s ANC
- deposition relationship facilitates the 22 linkage between ecos ys tem models that calculate an ecological indicator and the atmospheric 23Draft

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118 deposition of NOy

and SOx. On the other hand, there is no direct linear relationship between 1 deposition and pH. Whil e acid inputs from deposition lower pH, the relationship can be 2 extremely

nonlinear and there is no direct connection from a modeling or mass balance 3 perspective between the amount of deposition entering a s ys tem and pH. The term “mass 4 balance” underlies the basic formulation of any phy sical modeling construct, atmospheric or 5 aquatic s ys tems, and refers to the accounting of the flow of mass into a s ys tem, the 6 transformation to other forms, and the loss due to flow out of a s ys tem and other removal 7 processe s. ANC is a conserved propert y. This means that ANC in a water body can be 8 accounted for b y knowledge of how much ANC initially exists, how much flows in and is 9 deposited, and how much flows out. In contrast, hydrogen ion concentration in the water, the 10 basis for pH, is not a conserved propert y as its concentration is affected b y several factors such 11 as temperature, atmospheric pressure, mixing conditions of a water body, and the levels of other 12 several chemical species in the sy stem. The disadvantage o f pH lacking conservative properties 13 is that there is a very complex connection between changes in ambient air concentrations of NOy 14 and SOx and pH. 15 The discussion of basic water chemistry of natural sy stems in chapter 2 of the Policy 16 Assessment provides further details on wh y pH is not a conserved quantity and is subject to rapid 17 transient response behavior that makes it difficult to use as a reliable and functional ecological 18 indicator. The observed pH

- to - ANC relationship ( US EPA, 2011, figure 7
- 2 ) partiall y explains 19 the concern with pH responding too abruptl y. In the region where pH ranges roughly from 4.5 20 to 6 and is of greatest relevance to effects ( US EPA, 2011, figure 7

- 4 ), there clearly is more 21 sensitivity

of pH to changes in ANC in the ANC rang e from approximately 0 to 50 μeq/L . A 22 focus on this part of the ANC

- to - pH relationship shows that ANC associates well with pH in a 23Draft

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119 fairl y linear manner. However, the pH range from 4.5 to 6 also includes one of the very steepest 1 parts of the slope relati ng pH as a function of ANC , where ANC ranges down below 0 μeq/L, 2 which is subject to very rapid change in ANC, or deposition inputs. This part of the relationship 3 coincides with reduced levels of ANC and hence with reduced ability to neutralize acids and 4 moderate pH fluctuations. This response behavior can be extended to considering how pH 5 would change in response to deposition, or ambient concentrations, of NOy and SOx, which can 6 be viewed as “ANC

- like” inputs. 7 In summary , because ANC clearl y links both to biological effects of aquatic acidification 8 as well as to acidifying inputs of NOy and SOx deposition, the Policy Assessment concludes that 9 ANC is an appropriate ecological indicator for relating adverse aquatic ecosystem effects to 10 acidifying atmo spheric deposition of SOx and NOy , and is preferred to other potential indicators. 11 In reaching this conclusion, the Policy Assessment notes that in its review of the first draft Policy 12 Assessment, CASAC concluded that “information on levels of ANC protect ive to fish and other 13 aquatic biota has been well developed and presents probably the lowest level of uncertainty in 14 the entire methodology” (Russell and Samet, 2010a). In its more recent review of the second 15 draft Policy

Assessment, CASAC agreed “that ac id neutralizing capacit y is an appropriate 16 ecological measure for reflecting the effects of aquatic acidification” (Russell and Samet, 2010b; 17 p. 4). 18 2. Linking ANC to Deposition 19 There is evidence to support a quantified relationship between deposition of nitrogen and 20 sulfur and ANC. This relationship was anal yz ed in the REA for two case study areas, the 21 Adirondack and Shenandoah Mountains, based on time

- series modeling and observed trends. 22Draft

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120 In the REA anal ys is, long

- term trends in surface water nitrate, sulfate and ANC were modeled 1 using Model of Acidification of Groundwater in Catchment (MAGI C) for the two case study 2 areas. These data were used to compare recent surface water conditions in 2006 with 3 preindustrial conditions (i.e. preacidification 18 60). The results showed a marked increase in the 4 number of acid impacted lakes, characterized as a decrease in ANC levels, since the onset of 5 anthropogenic nitrogen and sulfur deposition, as discussed in chapter 2 of the Policy Assessment. 6 In the REA, mor e recent trends in ANC, over the period from 1990 to 2006, were assessed using 7 monitoring data collected at the two case study areas. In both case study areas, nitrate and 8 sulfate deposition decreased over this time period. In the Adirondack Mountains, this 9 corresponded to a decreased concentration of nitrate and sulfate in the surface waters and an 10 increase in ANC ( US EPA, 2009, section 4.2.4.2). In the Shenandoah Mountains, there was a 11 slight decrease in nitrate and sulfate concentration in surface waters corresponding to modest 12 increase in ANC from 50 ueq/Lin 1990 to 67 ueq/L in 2006 ( US EPA, 2009, section 13 4.2.4.3 , Appendix 4, and section 3.4). 14 In the REA, the quantified relationship between deposition and ANC was investigated 15 using ecos ys tem acidification models, also referred to as acid balance models or critical loads 16 models ( US EPA, 2011 , section 2 and US EPA, 2009, secti on 4 and Appendix 4). These models 17 quantify

the relationship between deposition of nitrogen and sulfur and the resulting ANC in 18 surface waters based on an ecos ys tem’s inherent generation of ANC and ability to neutralize 19 nitrogen deposition through biologi cal and ph ys ical processes. A critical load is defined as the 20 amount of acidifying atmospheric deposition of nitrogen and sulfur beyond which a target ANC 21 is not reached. Relatively high critical load values imply that an ecos ys tem can accommodate 22Draft

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121 great er deposition levels than lower critical loads for a specific target ANC level. Ecos ys tem 1 models that calculate critical loads form the basis for linking deposition to ANC. 2 As discussed in chapter 2 of the Policy Assessment, both dy

namic and steady

s tate 3 models calculate ANC as a function of ecos ys tem attributes and atmospheric nitrogen and sulfur 4 deposition, and can be used to calculate critical loads. Steady state models are time invariant 5 and reflect the long term consequences associated with an ecos ys tem reaching equilibrium under 6 a constant level of atmospheric deposition. Dy namic models are time variant and take into 7 account the time dependencies inherent in ecos ys tem hydrology , soil and biological processes. 8 Dy namic models like MAGI C can p rovide the time series response of ANC to deposition 9 whereas stead y state models provide a single ANC relationship to any fixed deposition level. 10 Dy namic models naturally

are more complex than steady state models as they

attempt to capture 11 as much of the fundamental biogeochemical processes as practicable, whereas steady state 12 models depend on far greater parameterization and generalization of processes that is afforded, 13 somewhat, b y not having to accounting for temporal variability. 14 The Policy

Assess ment notes that stead y state models are capable of addressing the 15 question of what does it take to reach and sustain a specific level of ANC
. Dy namic models are 16 also capable of addressing that question, but can also address the question of how long it tak es to 17 achieve that result. Dy namic models afford the ability for more comprehensive treatment of a 18 variet y of processes throughout the surface, soil and bedrock layers within an ecos ys tem. For 19 example, steady

state models treat sulfate as a mobile anion throughout the sy stem, meaning that 20 the sulfate that is deposited to a watershed enters the water column and is not influenced by soil 21 adsorption or cation exchange. Dy namic models can incorporate these time variant processes. 22 The use of a stead y state m odel treating sulfate as totally mobile does not necessarily conflict 23Draft

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122 with the possibility

of sulfate acting as a less than mobile ion at certain times. The steady state 1 assumption is premised on the long term behavior of sulfate which can undergo period s of net 2 adsorption followed b y periods of net desorption which can balance out over time. The Policy 3 Assessment recognizes that as the richness of the available data increases, in terms of parameters 4 and spatial resolution, the incorporation of dy namic m odeling approaches in the standard setting 5 process should become more feasible. In determining an appropriate modeling approach for the 6 development of a NAAQS in this review, the Policy Assessment considers both the relevance of 7 the question addressed as well as the ability to perform modeling that provides relevant 8 information for geographic areas across the country. 9 Dy

namic models require a large amount of catchment level

- specific data relative to 10 steady

state models. Because of the time invariant nature of stead y state models, the data 11 requirements that integrate across a broad spectrum of ecos ys tem processes is achievable and 12 available now at the national level. Water quality data to support stead y state models currently 13 exist for developing a natio nal data base for modeling nearl y 10,000 catchments in the 14 contiguous U.S. In contrast, the data needs to support dy namic models for national

- scale 15 anal ys es simply

are not available at this time. Further, the information provided by steady state 16 modeling would be sufficient to develop and analyze alternative NAAQS and the kind of 17 protection they

would afford. While it would be of interest to also obtain information about how 18 much time it would take for a target ANC level to be achieved, the absence of su ch information 19 does not preclude developing and evaluating alternative NAAQS using the AAI structure. Based 20 on the above considerations, the Policy Assessment concludes that at this time steady state 21 critical load modeling is an appropriate tool for linki ng long

- term ANC levels to atmospheric 22 deposition of nitrogen and sulfur for development of an AAI that has national applicability

. 23Draft

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123 A s tead y state model is used to define the critical load, which is the amount of 1 atmospheric deposition of nitrogen (N) and sulfur (S) beyond which a target ANC is not 2 achieved and sustained. 6 It is expressed as: 3 4 CLANClim (N + S) = ([BC] 0 *

- [ANC lim ]) Q + Neco (III

- 1) 5 6 Where: 7 CLANClim (N + S) is the critical load of deposition, with units of equivalent charge/(area

- time ); 8 [BC] 0 * is the natural contribution of base cations from weathering, soil processes and 9 preindustrial deposition, with units of equivalent charge/volume; 10 [ANC

lim ] is the target ANC value, with units of equivalent charge/volume; 11 Q is the catchment level runoff rate governed b y water mass balance and dominated by 12 precipitation, with units of distance/time; and 13 Neco is the amount of nitrogen deposition that is effectivel y neutralized by a variet y of biological 14 (e.g., nutrient uptake) and phy sical processe s, with units of equivalent charge/ (area

- time). 15 16 Equation III

- 1 is a modified expression that adopts the basic formulation of the steady 17 state models that are described in chapter 2 of the Policy Assessment. More detailed discussion 18 of the rationale, a ssumptions and derivation of equation III

- 1, as well as all of the equations in 19 this section, are included in Appendix B of the Policy Assessment. The equation simply reflects 20 the amount of deposition of nitrogen and sulfur from the atmosphere, CLANClim (N + S) , that is 21 associated with a sustainable long

- term ANC target, [ANC lim ] , given the capacity of the natural 22

6 This section discusses the linkages between deposition of nitrogen and sulfur and ANC. Section III.B.3 then discusses the link ages between atmospheric concentrations of NOy and SOx and deposition of nitrogen and sulfur

.Draft

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124 sy stem to generate ANC, [BC] 0 * , and the capacit y of the natural s ys tem to neutralize nitrogen 1 deposition, Neco . This expression of critical l oad is valid when nitrogen deposition is greater 2 than Neco. 7 The runoff rate, Q , allows for balancing mass in the two environmental mediums

– 3 atmosphere and catchment. This critical load expression can be focused on a single water s ys tem 4 or more broadl y. To extend applicability of the critical load expression (equation III

- 1) from the 5 catchment level to broader spatial areas, the terms Q r and CLr, are used, which are the runoff rate 6 and critical load, respectively , of the region over which all the atmosp heric terms in the equation 7 are defined. 8 In considering the contributions of SOx or NOy species to acidification, it is useful to 9 think of every

depositing nitrogen atom as supplying one equivalent charge unit and every sulfur 10 atom as depositing two char ge units. The Policy Assessment uses equivalent charge per volume 11 as a normalizing tool in place of the more familiar metrics such as mass or moles per volume. 12 This allows for a clearer explanation of many of the relationships between atmospheric and 13
eco sy stem processes that incorporate mass and volume unit conventions somewhat specific to 14 the environmental media of concern (e.g., m 3 for air and liter for liquid water). Equivalent 15 charge reflects the chemistry equilibrium fundamentals that assume electro neutrality , or 16 balancing charge where the sum of cations alway s equals the sum of anions. 17 As presented above, the terms S and N in the CLANClim (N + S) term broadly represent all 18 species of sulfur or nitrogen that can contribute to acidifying deposition. This follows 19 conventions used in the scientific literature that addresses critical loads, and it reflects all 20 possible acidifying contributions from any sulfur or nitrogen species. For all practical purposes, 21 S reflects SOx as described above, the sum o f sulfur dioxide gas and particulate sulfate. 22

7 Because Neco is only relevant to nitrogen deposition, in rare cases where Neco is greater than the total nitrogen deposition, the critical load would be defined only in terms of acidifying deposition of sulfur and the Neco term in equation III

- 1 wo uld be set to zero.Draft

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125 However, N in equation III

- 1 includes both oxidized forms, consistent with the ambient 1 indicator, NOy , in addition to the reduced nitrogen species, ammonia and ammonium ion, 2 referred to as NHx. NHx is inclu ded in the critical load formulation because it contributes to 3 potentially

acidifying nitrogen deposition. Consequently , from a mass balance or modeling 4 perspective, the form of the standard needs to account for NHx, as described below. 5 3. Linking Deposit ion to Ambient Air Indicators 6 The last major component of the form illustrated in Figure III

- 1 addresses the linkage 7 between deposition of nitrogen and sulfur and concentrations of the ambient air indicators, NOy 8 and SOx. To link ambient air concentrat ions with deposition, the Policy Assessment defines a 9 transference ratio, T, as the ratio of total wet and dry deposition to ambient concentration, 10 consistent with the area and time period over which the standard is defined. To express 11 deposition of NOy

and SOx in terms of NOy and SOx ambient concentrations, two transference 12 ratios were defined, where T SOx equals the ratio of the combined dry and wet deposition of SOx 13 to the ambient air concentration of SOx, and T NOy equals the ratio of the combined dry an d wet 14 deposition of NOy

to the ambient air concentration of NOy

. 15 As described in Chapter 7 of the Policy Assessment, reduced forms of nitrogen (NH X ) are 16 included in total nitrogen in the critical load equation, III

- 1. Reduced forms of nitrogen are 17 treat ed separately , as are NOy and SOx, and the transference ratios are applied. This results in 18 the following critical load expression that is defined explicitly in terms of the indicators NOy and 19 SOx: 20 21 CLANClim (N + S) = ([BC] 0 *

- [ANC lim ]) Q + Neco = [NOy

]T NO y + [SOx ]T SOx + NHx (III

- 2) 22 23Draft

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126 This is the same equation as III

- 1, with the deposition associated with the critical load translated 1 to deposition from ambient air concentrations via transference ratios. In addition, deposition of 2 reduced nitrogen, oxidize nitrogen and oxidized sulfur a re treated separatel y. 3 Transference ratios are a modeled construct, and therefore cannot be compared directl y to 4 measurable quantities. There is an analogy to deposition velocity , as a transference ratio is 5 basically

an aggregated weighted average of the deposition velocities of all contributing species 6 across dry

and wet deposition, and transference ratio units are expressed as distance/time. 7 However, wet deposition commonly is not interpreted as the product of a concentration times a 8 velocity . Direc t wet deposition observations are available which integrate all of the processes, 9 regardless of how well they may be understood, related to wet deposition into a measurable 10 quantity . There are reasonable analogies between the processes governing dry and w et 11 deposition, from a fundamental mass transfer perspective. In both cases there is a transfer of 12 mass between the dry ambient phase and another medium, either a surface or vegetation in the 13 case of dry

deposition, or a rain droplet or cloud in the case o f wet precipitation. The specific 14 thermody

namic properties and chemical/biological reactions that govern the transfer of dry mass 15 to plants or aqueous droplets differ, but either process can be based on conceptualizing the 16 product of a concentration, or co ncentration difference, times a mass transfer coefficient which is 17 analogous to the basic dry deposition model: dry deposition = concentration x velocity ( US 18 EPA, 2011, Appendix F ). 19 Transference ratios require estimates of wet deposition of NOy and SOx , dry deposition 20 of NOy

and SOx, and ambient air concentrations of NOy and SOx. Possible sources of 21 information include model estimates or a combination of model estimates and observations, 22 recognizing that dry deposition is a modeled quantity that can us e observed or modeled estimates 23Draft

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127 of concentration. The limited amount of NOy measurements in acid

- sensitive areas as well as 1 the combination of representative NOy , SO 2 and SO 4 observations generally preclude the use of 2 observations for development of a standard that is applicable nationall y. 3 The Policy

Assessment considers a blending of observations and models to take 4 advantage of their relative strengths; e.g., combining the NADP wet deposition observations, 5 modeled dry

deposition, and a mix of modeled an d observed concentrations, using the model for 6 those species not measured or measured with very sparse spatial coverage. A potential 7 disadvantage of mixing and matching observations and model estimates is to lose consistency 8 afforded b y using just modelin g alone. A modeling platform like CMAQ is based on adhering to 9 consistent treatment of mass conservation, by linking emission inputs with air concentrations and 10 concentrations to deposition. Inconsistencies from combining processes from different anal yt i cal 11 platforms increase the chance that mass (of nitrogen or sulfur) would unintentionally be 12 increased or decreased as the internal checking that assures mass conservation is lost. 13 Transference ratios incorporate a broad suite of atmospheric processes an d consequentl y 14 an anal yt ical approach that instills consistency in the linkage of these processes is preferable to 15 an approach lacking such inherent consistency . This contention does not mean that observations 16 alone, if available, could not be used, but s uggests that the inconsistencies in combining models 17 and observations for the purposes of developing transference ratios has the potential for creating 18 unintended artifacts. 19 While there is a reasonable conceptual basis for the concept of an aggregated deposition 20 velocity

referred to in the Policy

Assessment as a transference ratio, there is very limited ability 21 to compare observed and calculated ratios. This is because the deposition velocity is dependent 22 on individual species, and the mass transfer p rocesses of wet and dry removal, while 23Draft

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128 conceptuall y similar, are different. Consequently, there does not exist a meaningful approach to 1 measure such an aggregated or lumped parameter. Therefore, at this time, the evaluation of 2 transference ratios is base d on sensitivity studies, anal ys is of variabilit y, and comparisons with 3 other models, as described in Appendix F of the Policy Assessment. 4 As discussed in Appendix F, the interannual variability , as well as the sensitivity to 5 emission changes of roughly 50%, result s in changes of transference ratios of approximately 5
- 6 10%. Part of the reason for this inherent stability is due to the co

- dependence of concentration 7 and deposition. For example, as concentrations are reduced as a result of emissions reduc tions, 8 deposition in turn is reduced since deposition is a direct linear function of concentration leading 9 to negligible impact on the deposition

- to - concentration ratio. L ikewise, an overestimate of 10 concentration likely does not induce a bias in the trans ference ratio. While it is important to 11 continue to improve the model’s ability to match ambient concentrations in time and space, the 12 bias of a modeled estimate of concentration relative to observations does not necessaril y result in 13 a bias in a calculat ed transference ratio. In effect, this consideration of bias cancellation reduces 14 the sensitivity

of transference ratios to model uncertainties and affords increased confidence in 15 the stability

of these ratios. Based on the series of sensitivity and vari abilit y anal ys es, the Policy 16 Assessment concludes that the transference ratios are relativel y stable and provide a sound 17 metric for linking deposition and concentration in the form of the standard. 18 As discussed in the policy Assessment, transference ratio s are dependent on the platform 19 upon which they

are constructed. Comparisons of transference ratios constructed from different 20 modeling platforms do exhibit significant differences. While this divergence of results may be 21 explained by

a variet y of differ ences in process treatments, input fields and incommensurabilities 22 in species definitions and spatial configurations, it does suggest two very important conclusions. 23Draft

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129 First, the idea of using multiple platforms for different parts of the country may be pro blematic 1 as there does not exist a reliable approach to judge acceptance which is almost alway s based on 2 comparisons to observations. Second, since transference ratios are based on concentrations and 3 deposition, as the uncertainties in each of those compo nents are reduced, the relative uncertaint y 4 in the ratios also is reduced. This means that basic improvements in the model’s ability to 5 reproduce observed wet deposition and ambient concentration fields enhance the relative 6 confidence in the constructed transference ratios. Similarly , as in

- situ dry deposition flux 7 measurements become available that enable a more rigorous evaluation and diagnosis of 8 modeled dry

deposition processes, the expected improved treatment of dry deposition also would 9 increase con fidence in transference ratios. Finally , deposition is directl y related to ambient air 10 concentrations. Models like CMAQ rel y on the concentration

- to - deposition linkage to calculate 11 deposition, which is the foundation for broadl y based and robust assessme nts addressing 12 atmospheric deposition. In principle, the use of a modeled constructed transference ratio is based 13 on the same premise b y which we use models to estimate deposition in the first place. 14 The shortage of widel y available ambient air observati ons and the fact that estimates of 15 dry deposition requires modeling, collectively suggests that a unified modeling platform is the 16 best approach for constructing transference ratios. The Policy Assessment ( US EPA, 17 2011, section 2) considers CMAQ and other models, such as CAMx and the Canada’s AURAMS 18

- A Unified Regional Air

- quality Modeling S ys tem (Smy

the et al., 2008), and concludes that 19 CMAQ is the preferred modeling platform for constructing transference ratios for purposes of 20 developing a new secondary standard for consideration. This conclusion reflects the view that 21 for the purposes of defining transference ratios, a modeling platform should (1) be a multiple 22 pollutant model recognizing the m yriad of connections across pollutant categories that direct ly 23Draft

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130 and indirectl y impact nitrogen and sulfur characterization, (2) include the most comprehensive 1 scientific treatments of atmospheric processes that relate directly and indirectly

to characterizing 2 concentrations and deposition, (3) have an infrastructure capability that accommodates the 3 inclusion of improved scientific treatments of relevant processes and important input fields, and 4 (4) undergo frequent reviews regarding the adequacy of the underlying science as well as the 5 appropriateness in applications . The CMAQ platform exhibits all these characteristics. It has 6 been (and continues to be) extensively evaluated for several pollutant categories, is supported by 7 a central infrastructure of EPA scientists, with considerable interfacing with the scientifi c 8 research communities in academia and industry , whose mission is to improve and evaluate the 9 CMAQ platform. More directly , CMAQ, and its predecessor versions, has a long track record 10 going back to the NAPAP in the 1980’s of specific improvements in depo sition processes, which 11 are described in Appendix F of the Policy Assessment. 12 4. Aquatic Acidification Index 13 Having established the various expressions that link atmospheric deposition of nitrogen 14 and sulfur to ANC and the transference ratios that translate atmospheric concentrations to 15 deposition of nitrogen and sulfur, the Policy Assessment derives the following expression of 16 these linkages, which separates reduced forms of nitrogen, NHx, from oxidized forms: 17 18 ANCcalc = {ANClim + CL r /Q r }

– NHx/Q r - T NOy [NOy ]/ Q r - T SOx[SOx ]/ Qr (III

- 3) 19 20 Equation III

- 3 is the basic expression of the form of a standard that translates the 21 conceptual framework into an explicit expression that defines ANC as a function of the ambient 22 air indicators, NOy and SOx. T his equation calculates an expected ANC value based on ambient 23Draft

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131 concentrations of NOy

and SOx. The calculated ANC will differ from the target ANC (ANClim) 1 depending on how much the nitrogen and sulfur deposition associated with NOy, SOx, and NHx 2 differs fr om the critical load associated with just achieving the target ANC. 3 Based on equation III

- 3, the Policy Assessment defines an aquatic acidification index (AAI ) that 4 is more simply

stated using terms that highlight the ambient air indicators: 5 6 AAI = F1

– F2

– F3[NOy ]

– F4[SOx

] (III

- 4) 7 8 where the AAI represents the long term (or stead y state) ANC level associated with ambient air 9 concentrations of NOy and SOx. The factors F1 through F4 convey three attributes: a relative 10 measure of the ecos ys tem ’s ability to neutralize acids (F1), the acidifying potential of reduced 11 nitrogen deposition (F2), and the deposition

- to - concentration translators for NOy (F3) and SOx 12 (F4). 13 Specificall y: 14 F1 = ANClim + CL r /Q r ; 15 F2 = NHx/ Q r = NHx deposition divided by Q r ; 16 F3 = T NOy/ Q r ; T NOy is the transference ratio that converts ambient air concentrations of NOy to 17 deposition of NOy ; and 18 F4 = T SOx/ Q r ; T SOx is the transference ratio that converts ambient air concentrations of SOx to 19 deposition of SOx. 20 All of these factors include representative Q r to maintain unit (and mass) consistency between 21 the AAI and the terms on the right side of equation III

- 4. 22Draft

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132 The F1 factor incorporates an ecos ys tem’s ability to generate acid neutralizing capacit y 1 through base cation supply ([BC] * 0 ) and to neutralize acidifying nitrogen deposition through 2 Neco, both of which are incorporated in the CLterm. As noted above, because Neco can onl y 3 neutralize nitrogen deposition (oxidized or reduced) there may be rare cases where Neco exceeds 4 the combination of reduced and oxidized nitrogen deposition. Consequently , to ensure that the 5 AAI equation is applicable in all cases that may occur, equation III

- 4 is conditional on total 6 nitrogen deposition, {NHx + F3[NOy]}, being greater than Neco. In rare cases where Neco is 7 greater than {NHx + F3[NOy ]}, F2 are F3 would be set equal to 0 in the AAI equation. The 8 consequence of setting F2 and F3 to zero is simply to constrain the AAI calculation just to SOx, 9 as nitrogen would have no bearing on acidifying contributions in this case. 10 The Policy

Assessment concludes that equation III

- 4 ( US EPA, 2011, equation 7

- 12), 11 which defines the AAI, is ecologicall y relevant and appropriate for use as the form of a national 12 standard designed to provide protection for aquatic ecos ys tems from the effects association with 13 acidifying deposition associated with concentrations of oxides of nitrogen and sulfur in the 14 ambient air. This AAI equation does not, however, in itself, define the spatial areas over wh ich 15 the terms of the equation would apply . To specify values for factors F1 through F4, it is 16 necessary

to define spatial areas over which these factors are determined. Thus, it is necessary to 17 identify

an approach for spatially aggregating water bodies into ecologically meaningful regions 18 across the U.S., as discussed below. 19 5. Spatial Aggregation 20 As discussed in the Policy Assessment, one of the unique aspects of this form is the need 21 to consider the spatial areas over which values for the F factors in the AAI equation are 22 quantified. Ecos ys tems across the U.S. exhibit a wide range of geological, hydrological and 23Draft

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133 vegetation characteristics that influence greatl y the ecos ys tem parameters, Q, BC 0 * and Neco that 1 are incorporated in the AAI. Variations in ecos ys tem attributes naturall y lead to wide variabilit y 2 in the sensitivities of water bodies in the U.S. to acidification, as well as in the responsiveness of 3 water bodies to changes in acidifying deposition. Consequently , variations in ecosy stem 4 sensiti vity , and the uncertainties inherent in characterizing these variations, must be taken into 5 account in developing a national standard. In developing a secondary NAAQS to protect public 6 welfare, the focus of the Policy Assessment is on protecting sensitive populations of water 7 bodies, not on each individual water body, which is consistent with the Agency’s approach to 8 protecting public health through primary NAAQS that focus on susceptible populations, not on 9 each individual. 10 The approach used for defining ecologicall y relevant regions across the U.S. in the Policy 11 Assessment ( US EPA, 2011, section 7.2.5) is described below, along with approaches to 12 characterizing each region as acid sensitive or relatively non - acid sensitive. This 13 characterization facilitates a more detailed anal ys is and focus on those regions that are relativel y 14 more acid sensitive. This characterization is also used to avoid over

- protection in relatively non 15 acid sensitive regions, regions that would receive limited benefit from reductio ns in the 16 deposition of oxides of nitrogen and sulfur with respect to aquatic acidification effects. 17 Approaches to developing representative values for each of the terms in the AAI equation 18 (factors F1 through F4) for each ecologically relevant region for which sufficient data are 19 available are then discussed. These spatial aggregation approaches are generall y applicable to 20 the contiguous United States. The following discussion also addresses the development of 21 factors for data

- limited regions and specif ically for Hawaii, Alaska and the U.S. territories. 22Draft

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134 Stated more simply , this section discusses appropriate way s to divide the country into 1 ecologicall y relevant regions; to characterize each region as either acid sensitive or relativel y 2
non - acid sensitive ; and to determine values of factors F1 through F4 for each region, taking into 3 consideration the acid sensitivity of each region and the availability of relevant data. For each 4 such region, the AAI would be calculated based on the values of factors F1 through F4 specified 5 for that region. 6 In considering approaches to spatial aggregation, the Policy Assessment focuses on 7 methods that have been developed to define ecologically relevant regions, referred to as 8 ecoregions, which are meaningfull y related to the factors that are relevant to aquatic 9 acidification. As noted above, the Policy Assessment did not focus on looking at each individual 10 water bod y, nor did it focus on aggregating over the entire nation, which would preclude taking 11 into account the inher ent variability in atmospheric and ecological factors that fundamentall y 12 modify

the relationships that are central to the development of an ecologically relevant AAI. 13 Based on considering available classification schemes, the Policy Assessment concludes 14 that Omernik’s ecoregion classification is the most appropriate method to consider for the 15 purposes of this review. This classification offers several levels of spatial delineation, has 16 undergone an extensive scientific peer review process, and has explic itly been applied to 17 delineating acid sensitive areas within the U.S. Further, the Policy Assessment concludes that 18 ecoregion level III (Figure III
- 1 ) resolution, with 84 defined ecoregions in the contiguous U.S., 8 19 is the most appropriate level to consider for this purpose. The spatial resolution afforded b y 20 level III strikes an appropriate balance relative to the reasoning that supports conclusions on 21 indicators, as discussed above. The Policy Assessment conclude s that the most detailed level of 22 resolution (level IV) is not appropriate given the limited data availability to address nearl y 1000 23

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

.Draft

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135 subdivisons within that level and the currently evolving nature of level IV regions. Further, level 1 III ecoregions are pr eferred to level II in that level III ecoregions, but not level II ecoregions, are 2 largel y contiguous in space which allows for a more coherent development of information to 3 quantify

the AAI factors and to characterize the concentrations of NOy and SOx in the ambient 4 air within each ecoregion. 5 Appendix C of the Policy Assessment includes a description of each level III ecoregion. 6 The Policy

Assessment notes that the use of ecoregions is an appropriate spatial aggregation 7 scheme for this standard focused o n deposition

- related aquatic acidification effects, while man y 8 of the same ecoregion attributes may be applicable in subsequent NAAQS reviews that may 9 address other deposition

- related aquatic and terrestrial ecological effects. Because atmospheric 10 deposi tion is modified by ecos ys tem attributes, the ty pes of vegetation, soils, bedrock geology , 11 and topographic features that are the basis of this ecoregion classification approach also will 12 likely

be key attributes for other deposition

- related effects (e.g., terrestrial acidification, nutrient 13 enrichment) that link atmospheric concentrations to an aquatic or terrestrial ecological indicator. 14Draft

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13 6 Figure III - 1 . Omernik Ecoregion III areas ( http://www.epa.gov/we d/pages/ecoregions ). 1 2
a. Ecoregion Sensitivity 3 The Policy

Assessment used Omernik’s original alkalinity data ( US EPA, 2011, as section 4 2 ) and more recent ANC data to delineate two broad groupings of ecoregions: acid

- sensitive and 5 relativel y non - acid sensitive ecoregions. This delineation was made to facilitate greater focus on 6 those ecoregions with water bodies that generall y have greater acid sensitivity and to avoid over

7 protection in regions with generall y less sensitive water bodies. The approach used to delineate 8 acid

- sensitive and relatively non - acid sensitive regions included an initial numerical

- based 9 sorting scheme using ANC data , which categorized ecoregions with relatively high ANC values 10Draft

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137 as being relativel y non - acid sensitive. This initial delineation resulted in 29 of the 84 Omernik 1 ecoregions being categorized as acid sensitive. Subsequently , land use data were also conside red 2 to determine to what extent an ecoregion is of a relatively pristine and rural nature by quantifying 3 the degree to which active management practices related to development and agriculture occur in 4 each ecoregion. 5 The overall objective is to produc e a logical and practical grouping of ecoregions that 6 experience adverse conditions with respect to aquatic acidification and are likely to respond to 7 changes in concentrations of NOy and SOx in the ambient air and to the related deposition 8 levels. To ach ieve this goal, a two

- step process has been applied, first identifying acid sensitive 9 ecoregions based on water quality data alone, and second identifying among those acid

- sensitive 10 ecoregions those with highl y developed and managed areas. These highl y de veloped and 11 managed ecoregions are placed in a non

- acid sensitive category to avoid over protection beyond 12 what is requisite to protect public welfare. More specificall y, in determining an ecoregion’s acid 13 sensitivity

status in step 1, ANC data across the 84 ecoregions are sorted ( US EPA, 2011, section 14 7 ) to determine the number of water bodies within a region with ANC values suggestive of acid 15 sensitivity , so as to screen out regions with an overabundance of high ANC values. In reviewing 16 the ANC data, the Policy Assessment identified 29 ecoregions th at meet two criteria: (1) greater 17 than 5% of water bodies with data with ANC values less than 200 μeq/Land (2) greater than 1% 18 of water bodies with ANC values less than 100 μeq/L . In step 2, land use data were used to 19 identify

those acid sensitive ecoreg ions with significant managed areas that would not be 20 considered as having a relatively pristine and rural character. The percentage of the 21 combination of developed (residential, transportation, industrial and commercial) and 22 agricultural (croplands, pas tures, orchards, viney ards) land use was used as an indicator of 23Draft

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138 managed land use area. Forest cover was used as an indicator of non

- managed land use more 1 directly

reflecting the pristine quality of a region. Based on the 2006 National Land Cover Data 2 ba se (NLCD, http://www.epa.gov/mrlc/nlcd

- 2006.html ), acid sensitive ecoregions would meet 3 both of the following land use data criteria: percent of developed and agricultural area less than 4 20% combined with forested area greater than 50%. The combination of steps 1 and 2 identify 5 22 relativel y acid sensitive areas (Table III

- 1 and Figure III
- 2 ). 6 Consideration was also given to the use of naturally acidic conditions in defining 7 relativel y non - acid sensi tive areas. For example, several of the ecoregions located in plains near 8 the coast exhibit elevated dissolved organic carbon (DOC) levels, which is associated with 9 naturally

acidic conditions. DOC in surface waters is derived from a variety of weak organic 10 acid compounds generated from the natural availability and decomposition of organic matter 11 from biota. Consequently, high DOC is associated with “natural” acidity , with the implication 12 that a standard intended to protect against atmospheric contribut ions to acidity is not an area of 13 focus. The evidence suggests that several of the more highl y managed ecoregions in coastal or 14 near coastal transition zones are associated with relatively high DOC values, ty

picall y exceeding 15 on average 5 mg/l, compared t o other acid sensitive areas. Although there is sound logic to 16 interpret naturally

acidic areas as relatively non - acid sensitive, natural acidity indicators were not 17 explicitly

included in defining relativel y non - acid sensitive areas as there does not exi st a 18
consensus - based quantifiable scientific definition of natural acidity . Approaches to explicitly 19 define natural acidit y likely will be pursued in future reviews of the standard. 20Draft

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139 Table III - 1
. List of 22 Acid - Sensitive Areas Ecoregion nam e Ecoregion number Ridge and Valley 8.4.1 Northern Appalachian Plateau and Uplands 8.1.3 Piedmont 8.3.4 Western Allegheny Plateau 8.4.3 Southwestern Appalachians 8.4.9 Boston Mountains 8.4.6 Blue Ridge 8.4.4 Ouachita Mountains 8.4.8 Central Appalachians 8.4.2 Northern Lakes and Forests 5.2.1 Maine/New Brunswick Plains and Hills 8.1.8 North Central Appalachians 5.3.3 Northern Appalachian and Atlantic Maritime Highlands 5.3.1 Columbia Mountains/Northern Rockies 6.2.3 Middle Rockies 6.2.10 Wasatch and Uinta Mountains 6.2.13 North Cascades 6.2.5 Cascades 6.2.7 Southern Rockies 6.2.14 Sierra Nevada 6.2.12 Idaho Batholith 6.2.15 Canadian Rockies 6.2.4 1Draft

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140 1 Figure III - 2 . Acid

- sensitive ecoregions identified in grey fill (22 out of 84 ecoregions). 2 3 b. Representative Ecoregion

- specific Factors 4 Having concluded that the Omernik level III ecoregions are an appropriate approach to 5 spatial aggregation for the purpose of a standard to address deposition

- related aquatic 6 acidification effects, the Policy Asse ssment uses those ecoregions to define each of the factors in 7 the AAI equation. As discussed below, factors F1 through F4 in equation III

- 4 are defined for 8 each ecoregion by

specifyingecoregion

- specific values for each factor based on monitored or 9 modele d data that are representative of each ecoregion. 10Draft

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141 i. Factor F1 1 As discussed above, factor F1 reflects a relative measure of an ecos ys tem’s ability to 2 neutralize acidifying deposition, and is defined as: F1 = ANClim + CL r /Q r . The value of F1 for 3 each ecoregion would be based on a representative critical load for the ecoregion (CLr ) 4 associated with a single national target ANC level (ANClim, discussed below in section III.D), 5 as well as on a representative runoff rate (Q r ). To specify ecoregion

- specifi c values for the term 6 Q r , the Policy

Assessment used the median value of the distribution of Q values that are available 7 for water bodies within each ecoregion. To specify ecoregion

- specific representative values for 8 the term CLr in factor F1, a distribut ion 9 of calculated critical loads was created for the water 9 bodies in each ecoregion for which sufficient water quality and h ydrology

data are available. 10 10 The representative critical load was then defined to be a specific percentile of the distribution of 11 critical loads in the ecoregion. Thus, for example, using the 90 th percentile means that within an 12 ecoregion, 90 percent of the water bodies would be expected to have higher calculated critical 13 loads than the representative critical load. That is, if the representative critical load were to 14 occur across the ecoregion, 90 percent of the water bodies would be expected to achieve the 15 national ANC target or better. 16 The specific percentile selected as part of the definition of F1 is an important parameter 17 that directly

impacts the representative critical load specified for each ecoregion, and therefore 18 the degree of protectiveness of the standard. A higher percentile corresponds to a lower critical 19 load and, therefore, to lower allowable ambient air concentrations of NOy and SOx and related 20

9 T he distribution of critical loads w as based on CL values calculated with Neco at the lake level. Consideration could also be given to using a distribution of CLs without Neco and adding the ecoregion average Neco value to the nth percentile critical load. This would avoid cases where the lake

- level Neco value potentially could be greater than total nitrogen deposition. The CL at the lake level represents the CL for the lake to achieve the specified national target ANC value. 10 The Policy Assessment judge d the data to be sufficient for this purpose if data are available from more than 10 water bodies in an ecoregion.Draft

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142 deposition to achieve a target AAI level. In conjunction with the other terms in the AAI 1 equation, alternative forms can be appropriatel y characterized in part b y i dentifying a range of 2 alternative percentiles. The choice of an appropriate range of percentiles to consider for acid

3 sensitive and relativel y non - acid sensitive ecoregions, respectivel y, is discussed below. 4 ( a ) Acid

- sensitive Ecoregions 5 In identifying percentiles that are appropriate to consider for the purpose of calculating 6 factor F1 for ecoregions characterized as acid sensitive, the Policy Assessment concludes that it 7 is appropriate to focus on the upper part of the distribution of critical loads, so as to ensure that 8 the ecoregion would be represente d by relativel y more acid sensitive water bodies within the 9 ecoregion. Specifying factor F1 in this way would help to define a standard that would be 10 protective of the population of acid sensitive water bodies within an ecoregion, recognizing that 11
even ec oregions characterized as acid sensitive may contain a number of individual water bodies 12 that are not acid sensitive. The Policy Assessment recognizes that there is no basis for 13 independentl y evaluating the degree of protectiveness afforded b y an y specifi c percentile value, 14 since it is the combination of form and level, in conjunction with the indicator and averaging 15 time, which determine the degree of protectiveness of a standard. In light of this, the Policy 16 Assessment concludes that it is appropriate t o consider a range of percentiles, from well above 17 the 50 th percentile, or median, of the distribution to somewhat below the highest value. More 18 specificall y, the Policy Assessment concludes it is appropriate to consider percentiles in the 19 range of the 70 th to the 90 th percentile. This conclusion is based on the judgment that it would 20 not be appropriate to represent an ecoregion with the highest or near highest critical load, so as to 21 avoid potential extreme outliers that can be seen to exist at the extre me end of the data 22 distributions, which would not be representative of the population of acid sensitive water bodies 23Draft

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143 within the ecoregion. Also, in considering ecoregions that are inherently acid sensitive, it is 1 judged to be appropriate to limit the lowe r end of the range for consideration to the 70 th 2 percentile, a value well above the median of the distribution. 3 In considering this conclusion, the CASAC Panel noted that the data bases for calculating 4 critical loads within an ecoregion are not necess arily representative of all water bodies within an 5 ecoregion. That is, in many ecoregions the lake sampling design used in studies that generated 6 the relevant data may have focused on the relatively more sensitive water bodies within an 7 ecoregion (Russell and Samet, 2011 a). Consequently , a given percentile of the distribution of 8 calculated critical loads, based on sampled water bodies, may not be representative of that 9 percentile of all water bodies across an entire ecoregion. To the extent that the samp ling of water 10 bodies within an ecoregion was skewed toward the relativel y more sensitive water bodies, 11 selecting a given percentile from the distribution of available critical loads would result in a 12 somewhat higher percentile of all water bodies within th at ecoregion having a higher calculated 13 critical load than the representative critical load value. Thus, the extent to which study sampling 14 designs have resulted in skewed distributions of calculated critical loads is an uncertaint y that is 15 appropriate to consider in selecting a percentile for the purpose of defining the factor F1 in the 16 AAI equation. 17 ( b ) Non

- acid sensitive Ecoregions 18 With regard to identifying percentiles that are appropriate to consider for the purpose of 19 calculating factor F1 for ecor egions characterized as relativel y non - acid sensitive, the Policy 20 Assessment recognizes that while such ecoregions are generall y less sensitive to acidifying 21 deposition from oxides of nitrogen and sulfur, they may contain a number of water bodies that 22
are acid sensitive This category includes ecoregions that are well protected from acidification 23Draft

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144 effects due to natural production of base cations and high ANC levels, as well as naturall y acidic 1 sy stems with limited base cation production and consequently ve ry low critical loads. Therefore, 2 the use of a critical load that would be associated with a highl y sensitive water bod y in a 3 naturally

acidic s ys tem would impose a high degree of relative protection in terms of allowable 4 ambient air concentrations of oxi des of nitrogen and sulfur and related deposition, while 5 potentially

affording little or no public welfare benefit from attempting to improve a naturally 6 acidic s ys tem. 7 Based on these considerations, the Policy Assessment concludes it is appropriate to 8
co nsider the use of a range of percentiles that extends lower than the range identified above for 9 acid

- sensitive ecoregions. Consideration of a lower percentile would avoid representing a 10 relativel y non - acid sensitive ecoregion by a critical load associated with relatively more acid

11 sensitive water bodies. In particular, the Policy Assessment concludes it is appropriate to focus 12 on the median or 50 th percentile of the distribution of critical loads so as to avoid over

- protection 13 in such ecoregions. Recogn izing that relativel y non - acid sensitive ecoregions generall y are not 14 sampled to the extent that acid

- sensitive ecoregions are, it also is appropriate to consider using 15 the median critical load of all relativel y non - acid sensitive ecoregions for each such ecoregion. 16 ii. Factor F2 17 As discussed above, factor F2 is the amount of reduced nitrogen deposition within an 18 ecoregion, including the deposition of both ammonia gas and ammonium ion, and is defined as: 19 F2 = NHx/ Q r . The Policy Assessment calculated the representative runoff rate, Q r , using a 20 similar approach as noted above for factor F1; i.e., the median value of the distribution of Q 21 values that are available for water bodies within each ecoregion. In the Policy Assessment, 2005 22 CMAQ model simulations o ver 12
- km grids are used to calculate an average value of NHx for 23Draft

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145 each ecoregion. The NHx term is based on annual average model outputs for each grid cell, 1 which are spatially

averaged across all the grid cells contained in each ecoregion to calculate a 2
re presentative annual average value for each ecoregion. The Policy Assessment concludes that 3 this approach of using spatially averaged values is appropriate for modeling, largel y due to the 4 relativel y rapid mixing of air masses that t ypicall y results in relativel y homogeneous air quality 5 patterns for regionall y dispersed pollutants. In addition, there is greater confidence in using 6 spatially

averaged modeled atmospheric fields than in using modeled point

- specific fields. 7 This averaging approach is also used for the air concentration and deposition terms in 8 factors F3 and F4, as discussed below. The Policy Assessment notes that modeled NHx 9 deposition exhibits greater spatial variability than the other modeled terms in factors F3 and F4. 10 Recognizing this gre ater variability , the Policy Assessment concludes that it would be 11 appropriate to consider alternative approaches to specifying the value of NHx. One such 12 approach might involve the use of more localized and/or contemporaneous modeling in areas 13 where this term is likel y to be particularly variable and important. 14 iii. Factors F3 and F4 15 As discussed above, factors F3 and F4 are the ratios that relate ambient air concentrations 16 of NOy

and SOx to the associated deposition, and are defined as follow: F3 = T NOy / Q r and F4 17 = T SOx/ Q r . T NOy is the transference ratio that converts ambient air concentrations of NOy to 18 deposition of NOy

and T SOx is the transference ratio that converts ambient air concentrations of 19 SOx to deposition of SOx. The representative r unoff rate, Q r , is calculated as for factors F1 and 20 F2. The transference ratios are based on the 2005 CMAQ simulations, using average values for 21 each ecoregion, as noted above for factor F2. More specifically , the transference ratios are 22 calculated as th e annual deposition of NOy or SOx spatially

averaged across the ecoregion and 23Draft

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146 divided by

the annual ambient air concentration of NOy or SOx, respectively , spatiall y averaged 1 across the ecoregion. 2
c. Factors in Data

- limited Ecoregions 3 As discussed above in section III.B.5.a, in the Policy Assessment the initial delineation of 4 acid

- sensitive and relatively non - acid sensitive ecoregions was based on available ANC and 5 alkalinity

data. Areas not meeting the ANC criteria described above are categorized as 6
r elativel y non - acid sensitive. The development of a reasonable distribution of critical loads for 7 water bodies within an ecoregion for the purpose of identifying the representative critical load 8 requires additional data, including more specific water quali ty data for major cations and anions. 9 This means that the water bodies that can be used to develop a distribution of critical loads is 10 generall y a subset of those water bodies for which ANC data are available Consequently , there 11 are certain ecoregions w ith sparse data that are not suitable for developing a distribution of 12 critical loads. 13 As noted above, the Policy Assessment judges that it is not appropriate to develop such 14 distributions based on data from less than ten water bodies within an ecoregion. Twelve such 15 ecoregions, which included only relativel y non - acid sensitive ecoregions, were characterized as 16 being data

- limited. For these ecoregions, the Policy Assessment considered alternative 17 approaches to specifying values for the terms CLr and Q r f or the purpose of determining values 18 for each of the factors in the AAI equation. For these data

- limited ecoregions, the Policy 19 Assessment judges that it is appropriate to use the median values of CLr and Q r from the 20 distributions of these terms for all o ther relatively non

- acid sensitive ecoregions, rather than 21 attempting to use severely limited data to develop a value for these terms based solel y on data 22 from such an ecoregion. Further, consideration could be given to using a single national default 23Draft

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147 val ue for all relativel y non - acid sensitive ecoregions. The Policy Assessment notes that this data 1 limitation is not a concern in specifying values for the other terms in the AAI equation for such 2 ecoregions, since those terms are based on data from the 2005 CMAQ model simulation, which 3 covers all ecoregions across the contiguous U.S. 4 d. Application to Hawaii, Alaska, and the U.S. Territories 5 The above methods for specifying ecoregion

- specific values for the factors in the AAI 6 equation appl y to those ecoregi ons within the contiguous U.S. For areas outside the continental 7 U.S., including Hawaii, Alaska, and the U.S. Territories, there is currentl y a lack of available 8 data to characterize the sensitivity of such areas, as well as a lack of water body

- specific data and 9 CMAQ
- ty

pe modeling to specify

values for the F1 through F4 factors. Thus, the Policy 10 Assessment has considered possible alternative approaches to specifying values for factors F1 11 through F4 in the AAI equation for these areas. 12 One such approach could be to specify area - specific values for the factors based on 13 values derived for ecoregions with similar acid sensitivities, to the extent that relevant 14 information can be obtained to determine such similarities. Such an approach would involve 15 conduc ting an analy sis to characterize similarities in relevant ecological attributes between 16 ecoregions in the contiguous U.S. and these areas outside the contiguous U.S. so as to determine 17 the appropriateness of utilizing ecoregion

- specific values for the CL r and Q r terms from one or 18 more ecoregions within the contiguous U.S. This approach would also involve conducting 19 additional air quality modeling for these area that are outside the geographical scope of the 20 currentl y available CMAQ model simulations, so as to develop the other information necessary 21 to specify

values for factors F2 through F4 for these areas. 22Draft

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148 A second approach could rely on future data collection efforts to establish relevant 1 ecological data within these areas that, together with additional air quality modeling, could be 2 used to specify

area - specific values for factors F1 through F4. Until such time as relevant data 3 become available, these areas could be treated the same as data

- limited ecoregions in the 4 contiguous U.S. that are relatively non - acid sensitive. 5 The Policy

Assessment concludes that either approach would introduce substantial 6 uncertainties that arise from attempting to extrapolate values based on similarity assumptions or 7 arbitraril y assigning values for factors in the AAI equation that would be applicable to these 8 areas outside the contiguous U.S. In light of such uncertainties, the Policy Assessment 9 concludes that it would also be appropriate to consider rel yi ng on the existing NO 2 and SO 2 10 secondary

standards in these areas f or protection of any

potential direct or deposition

- related 11 ecological effects that may be associated with the presence of oxides of nitrogen and sulfur in the 12 ambient air. The Policy Assessment concludes that rel yi ng on existing secondary standards in 13 th ese areas is preferable to using a highl y uncertain approach to allow for the application of a 14 new standard based on the AAI in the absence of relevant area

- specific data. 15 6. Summary of the AAI Form 16 With regard to the form of a multi

- pollutant air quality standard to address deposition

17 related aquatic acidification effects, the Policy Assessment concludes that consideration should 18 be given to an ecologically relevant form that characterizes the relationships between the 19 ambient air indicators for oxides of nitrogen and sulfur, the related deposition of nitrogen and 20 sulfur, and the associated aquatic acidification effects in terms of a relevant ecological indicator. 21 Based on the available information and assessments, consideration should be given to using a cid 22 neutralizing capacit y (ANC) as the most appropriate ecological indicator for this purpose, in that 23Draft

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149 it provides the most stable metric that is highl y associated with the water quality properties that 1 are directl y responsible for the principal adverse ef fects associated with aquatic acidification: 2 fish mortality

and reduced aquatic species diversity

. 3 The Policy

Assessment developed such a form, termed an aquatic acidification index 4 (AAI), using a simple equation to calculate an AAI value in terms of the ambient air indicators of 5 oxides and nitrogen and sulfur and the relevant ecological and atmospheric factors that modify 6 the relationships between the ambient air indicators and ANC. Recognizing the spatial 7 variability

of such factors across the U.S., th e Policy Assessment concludes it is appropriate to 8 divide the country

into ecologicall y relevant regions, characterized as acid

- sensitive or relativel y 9
non - acid

- sensitive, and specify the value of each of the factors in the AAI equation for each such 10 regio n. Omernik ecoregions, level III, are identified as the appropriate set of regions over which 11 to define the AAI. There are 84 such ecoregions that cover the continental U.S. This set of 12 ecoregions is based on grouping a variet y of vegetation, geological , and h ydrological attributes 13 that are directly

relevant to aquatic acidification assessments and that allow for a practical 14 application of an aquatic acidification standard on a national scale. 15 The Policy

Assessment defines AAIby the following equation: AAI = F1

– F2

– 16 F3[NO y ]

– F4[SO x ]. Factors F1 through F4 would be defined for each ecoregion b y specifying 17 ecoregion

- specific values for each factor based on monitored or modeled data that are 18 representative of each ecoregion. T he F1 factor is also defined b y a target ANC value. More 19 specificall y: 20 (1) F1 reflects a relative measure of an ecos ys tem’s ability to neutralize acidifying 21 deposition. The value of F1 for each ecoregion would be based on a representative 22 critical load for the ecoregion associated with a single national target ANC level, as well 23Draft

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150 as on a representative runoff rate. The representative runoff rate, which is also used in 1 specifying values for the other factors, would be the median value of the distributions of 2 runoff rates within the ecoregion. The representative critical load would be derived from 3 a distribution of critical loads calculated for each water body in the ecoregion for which 4 sufficient water qualit y and hydrology data are ava ilable. The representative critical load 5 would be defined b y s electing a specific percentile of the distribution. 6 In identifying a range of percentiles that are appropriate to consider for this 7 purpose, regions categorized as acid sensitive were consi dered separatel y from regions 8 categorized as relativel y non - acid sensitive. For acid sensitive regions, the Policy 9 Assessment concludes that consideration should be given to selecting a percentile from 10 within the range of the 70 th to the 90 th percentil e. The lower end of this range was 11 selected to be appreciably above the median value so as to ensure that the critical load 12 would be representative of the population of relatively more acid sensitive water bodies 13 within the region, while the upper end was selected to avoid the use of a critical load 14 from the extreme tail of the distribution which is subject to a high degree of variability 15 and potential outliers. For relativel y non - acid sensitive regions, the Policy Assessment 16 concludes that consider ation should be given to selecting the 50 th percentile to best 17 represent the distribution of water bodies within such a region, or alternatively to using 18 the median critical load of all relativel y non - acid sensitive areas, recognizing that such 19
areas ar e far less frequently evaluated than acid sensitive areas. Using either of these 20 approaches would avoid characterizing a generally non - acid

- sensitive region with a 21 critical load that is representative of relatively acid sensitive water bodies that may e xist 22 within a generall y non - acid sensitive region. 23Draft

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151 (2) F2 reflects the deposition of reduced nitrogen. Consideration should be given to 1 specifying the value of F2 for each region based on the averaged modeled value across 2 the region, using national CMAQ modeling that has been conducted b y EPA. 3 Consideration could also be given to alternative approaches to specifying this value, such 4 as the use of more localized and/or contemporaneous modeling in areas where this term is 5 likely

to be particularl y variabl e and important. 6 (3) F3 and F4 reflect transference ratios that convert ambient air concentrations of NO y 7 and SO x , respectivel y, into related deposition of nitrogen and sulfur. Consideration 8 should be given to specifyingthe values for F3 and F4 for eac h region based on CMAQ 9 modeling results averaged across the region. We conclude that specifying the values or 10 the transference rations based on CMAQ modeling results alone is preferred to an 11 alternative approach that combines CMAQ model estimates with obs ervational data. 12 (4) The terms [NO y ] and [SO x ] reflect ambient air concentrations measured at 13 monitoring sites within each region. 14 Using the equation, a value of AAI can be calculated for an y measured values of ambient 15 NOy

and SOx. For such a NAAQS, the Administrator would set a single, national value for the 16 level of the AAI used to determine achievement of the NAAQS, as discussed below in section 17 III.D. The ecoregion

- specific values for factors F1 through F4 would be specified by EPA based 18 on the most recent data and CMAQ model simulations, and codified as part of such a standard. 19 These factors would be reviewed and updated as appropriate in the context of each periodic 20 review of the NAAQS. 21 The Policy

Assessment developed specific F factors for each ecoregion based on the 22 approach discussed above, using alternative percentiles and alternative national target ANC 23Draft

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152 levels. The results of this analy sis for ecoregions characterized as acid sensitive are presented in 1 Table 7

- 1a
- d in the Policy Assessment. 2 C. Averaging Time 3 As discussed in section 7.3 of the Policy Assessment, aquatic acidification can occur over 4 both long

- and short

- term timescales. L ong - term cumulative deposition of nitrogen and sulfur is 5 reflected in the chronic acid

- base balance of surface waters as indicated by measured annual 6 ANC levels. Similarly , the use of stead y state critical load modeling, which generates critical 7 loads in terms of annual cumulative deposition of nitrogen and sulfur, means that the focus of 8 ecological eff

ects studies based on critical loads is on the long

- term equilibrium status of water 9 quality

in aquatic ecos ys tems. Much of the evidence of adverse ecological effects associated 10 with aquatic acidification, as discussed above in section II.A, is associated with chronicall y low 11 ANC levels. Protection against a chronic ANC level that is too low is provided by reducing 12 overall annual average deposition levels for nitrogen and sulfur. 13 Reflecting this focus on long

- term acidifying deposition, the Policy Assess ment 14 developed the AAI that links ambient air indicators to deposition

- related ecological effects, in 15 terms of several factors, F1 through F4. As discussed above, these factors are all calculated as 16 annual average values, whether based on water quality an d hydrology

data or on CMAQ model 17 simulations. In the context of a standard defined in terms of the AAI, the Policy Assessment 18 concludes that it is appropriate to consider the same annual averaging time for the ambient air 19 indicators as is used for the fa ctors in the AAI equation. 20 We also recognize that short

- term (i.e., hours or day s) episodic changes in water 21 chemistry , often due to changes in the hydrologic flow paths, can have important biological 22 effects in aquatic ecos ys tems. Such short

- term change s in water chemistry are termed “episodic 23Draft

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153 acidification.” Some streams may have chronic or base flow chemistry that is generall y healthy 1 for aquatic biota, but may be subject to occasional acidic episodes with potentially lethal 2 consequences. Thus, short

- term episodic ecological effects can occur even in the absence of 3 long

- term chronic acidification effects. 4 Episodic declines in pH and ANC are nearl y ubiquitous in drainage waters throughout the 5 eastern U.S. Episodic acidification can result from severa l mechanisms related to changes in 6 hydrologic flow paths. For example, snow can store nitrogen deposited throughout the winter 7 and snowmelt can then release this stored nitrogen, together with nitrogen derived from 8 nitrification in the soil itself, in a p ulse that leads to episodic acidification in the absence of 9 increased deposition during the actual episodic acidification event. The Policy Assessment notes 10 that inputs of nitrogen and sulfur from snowpack and atmospheric deposition largel y c ycl e 11 through soil. As a result, short

- term direct deposition inputs are not necessarily important in 12 episodic acidification. Thus, as noted in chapter 3 of the I SA, protection against episodic acidity 13 events can be achieved by establishing a higher chronic ANC level. 14 Taken together, the above considerations support the conclusion that it is appropriate to 15 consider the use of a long

- term average for the ambient air indicators NOy and SOx for an 16 aquatic acidification standard defined in terms of the AAI. The use of an annual averaging time 17 for NOy

and SOx concentrations would be appropriate to provide protection against low chronic 18 ANC levels, which in turn would protect against both long

- term acidification and acute acidic 19 episodes. 20 The Policy

Assessment has also con sidered interannual variability in both ambient air 21 quality

and in precipitation, which is directl y related to the deposition of oxides of nitrogen and 22 sulfur from the ambient air. While ambient air concentrations show ye ar - to - ye ar variability , 23Draft

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154 often the year - to - ye ar variability in precipitation is considerabl y greater, given the highl y 1 stochastic nature of precipitation. The use of multiple y ears over which annual averages are 2 determined would dampen the effects of interannual variability in both air qua lity

and 3 precipitation. For the ambient air indicators, the use of multiple

- ye ar averages would also add 4 stability

to calculations used to judge whether an area meets a standard defined in terms of the 5 AAI. Consequentl y, the Policy Assessment concludes that an annual averaging time based on 6 the average of each year over a consecutive 3 to 5 year period is appropriate to consider for the 7 ambient air indicators NOy and SOx. In reaching this conclusion, the Policy Assessment notes 8 that in its comments on th e second draft Policy Assessment, CASAC agreed that a 3 to 5 y ear 9 averaging time was appropriate to consider (Russell and Samet, 2010b). 10 D. Level 11 As discussed above, the Policy Assessment concludes that ANC is the ecological 12 indicator best suited to refl ect the sensitivity of aquatic ecos ys tems to acidifying deposition from 13 oxides of nitrogen and sulfur in the ambient air. ANC is an indicator of the aquatic acidification 14 expected to occur given the natural buffering capacity of an ecos ys tem and the loadi ngs of 15 nitrogen and sulfur resulting from atmospheric deposition. Thus, the Policy Assessment 16 developed a new standard for aquatic acidification that is based on the use of chronic ANC as the 17 ecological indicator as a component in the AAI. 18 The level of the standard would be defined in terms of a single, national value of the 19 AAI. The standard would be met at a monitoring site when the multi

- ye ar average of the 20 calculated annual values of the AAI was equal to or above the specified level of the standard. 11 21 The annual values of the AAI would be calculated based on the AAI equation using the assigned 22

11 Unlike other NAAQS, where the standard is met when the relevant value is at or below the level of the standard since a lower standard level is more protective, in this case a higher standard level is more protective.Draft

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155 ecoregion

- specific values for factors F1 through F4 and monitored annual average NOy and SOx 1 concentrations. Since the AAI equation is based on chronic ANC as the ecological indicator, the 2 level chosen for the standard would reflect a target chronic ANC value. As noted above, the 3 assigned F factors for each ecoregion would be determined b y EPA in the rulemaking to set the 4 NAAQS, based on water quality and h ydr ology

data, CMAQ modeling, the selected percentile 5 that is used to identify a representative critical load within the ecoregion, and the selected level 6 of the standard. The combination of the form of the standard, discussed above in section III.B, 7 defined by

the AAI equation and the assigned values of the F factors in the equation, other 8 elements of the standard including the ambient air indicators (section III.A) and their averaging 9 time (section III.C), and the level of the standard determines the allowa ble levels of NOy

and 10 SOx in the ambient air within each ecoregion. All of the elements of the standard together 11 determine the degree of protection from adverse aquatic acidification effects associated with 12 oxides of nitrogen and sulfur in the ambient air . The level of the standard play s a central role in 13 determining the degree of protection provided and is discussed below. 14 The Policy

Assessment focuses primaril y on information that relates degrees of 15 biological impairment associated with adverse ecologi cal effects to aquatic ecos ys tems to 16 alternative levels of ANC in reaching conclusions regarding the range of target ANC levels that 17 is appropriate to consider for the level of the standard. The Policy Assessment develops the 18 rationale for identifying a r ange of target ANC levels that is appropriate to consider by 19 addressing questions related to the following areas: (1) associations between ANC and pH levels 20 to provide an initial bounding for the range of ANC values to be considered; (2) evidence that 21 al lows for the delineation of specific ANC ranges associated with varying degrees of severity of 22 biological impairment ecological effects; (3) the role of ANC in affording protection against 23Draft

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156 episodic acidity ; (4) implications of the time lag response of ANC to changes in deposition; (5) 1 past and current examples of target ANC values applied in environmental management 2 practices; and (6) data linking public welfare benefits and ANC. 3 1. Association Between pH Levels and Target ANC Levels 4 As discussed above in section II.A and more full y in chapter 3 of the Policy Assessment, 5 specific levels of ANC are associated with differing levels of risk of biological impairment in 6 aquatic ecosy stems, with higher levels of ANC resulting in lower risk of ecosy stem impact s, and 7 lower ANC levels resulting in risk of both higher intensity of impacts and a broader set of 8 impacts. While ANC is not the causal agent determining biological effects in aquatic ecos ys tem, 9 it is a useful metric for determining the level at which a water bod y is protected against risks of 10 acidification. There is a direct correlation between ANC and pH levels which, along with 11 dissolved aluminum, are more closel y linked to the biological causes of ecosystem response to 12 acidification. 13 Because there is a direct correlation between ANC and pH levels, the selection of target 14 ANC levels is informed in part through information on effects of pH as well as direct studies of 15 effects related to ANC. Levels of pH are closely associated with ANC in the pH ran ge of 16 approximately

4.5 to 7. Within this range, higher ANC levels are associated with higher pH 17 levels. At a pH level of approximately 4.5, further reductions in ANC generally do not correlate 18 with pH, as pH levels remain at approximately 4.5 while ANC values fall substantially

. 19 Similarly , at a pH value of approximately 7, ANC values can continue to increase with no 20 corresponding increase in pH. As pH is the primary causal indicator of effects related to aquatic 21 acidification, this suggests that ANC va lues below approximately

- 50 μeq/L (the apparent point 22 in the relationship between pH and ANC where pH reaches a minimum) are not likely to result in 23Draft

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157 further damage. In addition, ANC values around and above approximately 100 μeq/L(the 1 apparent region in the relationship where pH reaches a maximum) are not likely to confer 2 additional protection. As a result, the initial focus in the Policy Assessment was on target ANC 3 values in the range of

- 50 to 100 μeq/L

. 4 2. ANC Levels Related to Effects on Aquatic Eco systems 5 As discussed above in section II.A, the number of fish species present in a water bod y has 6 been shown to be positively correlated with the ANC level in the water, with higher values 7 supporting a greater richness and diversit y of fish species. T he diversity and distribution of 8 phy to - zooplankton communities also are positively correlated with ANC. 9 A summary

of effects related to ANC ranges is shown above in Table II

- 1. Within the 10 ANC range from approximately

- 50 to 100 μeq/L, linear and sigmoid al relationships are 11 observed between ANC and ecos ys tem effects. On average, fish species richness is lower b y one 12 fish species for every 21 μeq/Ldecrease in ANC in Shenandoah National Park streams (I SA, 13 section 3.2.3.4). As shown in Table II

- 1, ANC lev els have been grouped into five categories 14 related to expected ecological effects, including categories of acute concern ( <0 μeq/L), severe 15 concern (0

- 20 μeq/L), elevated concern (20

- 50 μeq/L ), moderate concern (50

- 100 μeq/L), and 16 low concern (>100 μeq/L) . This categorization is supported by a large body

of research 17 completed throughout the eastern U.S. (Sullivan et al., 2006). 18 Water bodies with ANC values less than or equal to 0 μeq/Lat based flow are chronicall y 19 acidic. Such ANC levels can lead to complete loss of species and major changes in the ability of 20 water bodies to support diverse biota, especiall y in water bodies that are highly sensitive to 21 episodic acidification. Based on the above considerations, the Policy Assessment has focused on 22 target ANC levels no lower than 0 μeq/L. 23Draft

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158 As discussed in the Policy Assessment, biota generally are not harmed when ANC values 1 are >100 μeq/L , due to the low probability that pH levels will be below 7. In the Adirondacks, 2 the number of fish species also peaks at ANC values >100 μeq/L. This suggests that at ANC 3 levels greater than 100 μeq/L , little risk from acidification exists in many aquatic ecos ys tems. At 4 ANC levels below 100 μeq/L , overall health of aquatic communities can be maintained, although 5 fish fitness and community diversit y begin to decline . At ANC levels ranging from 100 down to 6 50 μeq/L , there is increasing likelihood that the fitness of sensitive species (e.g., brook trout, 7 zooplankton) will begin to decline. When ANC concentrations are below 50 μeq/L , the 8 probability

of acidification increases substantially , and negative effects on aquatic biota are 9 observed, including large reductions in diversity of fish species and changes in the health of fish 10 populations, affecting reproductive ability and fitness, especiall y in water bodies that are affected 11 by episodic acidification. While there is evidence that ANC levels above 50 can confer 12 additional protection from adverse ecological effects associated with aquatic acidification in 13 some sensitive e cos ys tems, the expectation that such incremental protection from adverse effects 14 will continue up to an ANC level of 100 is substantially reduced. The Policy

Assessment 15 concludes that the above considerations support a focus on target ANC levels up to a l evel 16 greater than 50 μeq/L but below 100 μeq/L , such as up to a level of 75 μeq/L. 17 In considering the available scientific evidence, as summarized here and discussed in 18 more detail in the ISA and REA, in its review of the second draft Policy Assessment, C ASAC 19 expressed the following views about the range of biological responses that corresponds to this 20 range of ANC levels (i.e., 0

- 100 μeq/L): 21 There will likel y be biological effects of acidification at higher ANC values 22 within this range, and there are relatively insensitive organisms that are not 23 impacted at ANC values at the low end of this range. Adverse effects of 24 acidification on aquatic biota are fairl y certain at the low end of this range of 25Draft

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159 ANC and incremental benefits of shifting waters to higher A NC become 1 more uncertain at higher ANC levels. There is substantial confidence that 2 there are adverse effects at ANC levels below 20 μeq/L , and reasonable 3 confidence that there are adverse effects below 50 μeq/L . Levels of 50 4 μeq/Land higher would provi de additional protection, but the Panel has less 5 confidence in the significance of the incremental benefits as the level 6 increases above 50 μeq/L. (Russell and Samet, 2010b) 7 8 The Policy

Assessment concludes that the above considerations, including the vi ews of 9 CASAC, provide support for focusing on target ANC levels in the range of 20 to 75 μeq/L

. 10 3. Consideration of Episodic Acidity 11 As discussed in the Policy Assessment, across the broad range of ANC values from 0 to 12 100 μeq/L , ANC affords protection against the likelihood of decreased pH (and associated 13 increases in Al) during long or short periods. In general, the higher the ANC wi thin this range, 14 the lower the probability of reaching low pH levels where direct effects such as increased fish 15 mortality

occur, as shown in Table 3

- 1 of the Policy Assessment. Accordingl y, greater 16 protection would be achieved b y target chronic ANC value s set high enough to avoid pH 17 depression to levels associated with elevated risk. 18 The specific relationship between ANC and the probability of reaching pH levels of 19 elevated risk varies b y water bod y and fish species. ANC levels below 20 μeq/Lare gen erall y 20 associated with high probability of low pH, leading to death or loss of fitness of biota that are 21 sensitive to acidification ( US EPA, 2008 , section 5.2.2.1; US EPA, 2009 , section 5.2.1.2). At 22 these levels, during episodes of high acidifying deposit ion, brook trout populations may 23 experience lethal effects. In addition, the diversity and distribution of zooplankton communities 24 decline sharpl y at ANC levels below 20 μeq/L. Overall, there is little uncertainty that significant 25 effects on aquatic biot a are occurring at ANC levels below 20 μeq/L

. 26Draft

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160 It is clear that at ANC levels approaching 0 μeq/L (Table II

- 1), there is significant 1 impairment of sensitive aquatic ecos ys tems with almost complete loss of fish species. Avoiding 2 ANC levels approaching 0 μeq/Lis particularl y relevant to episodic spikes in acidity that occur 3 during periods of rapid snow melt and during and after major precipitation events. Since the 4 ANC range considered in the Policy Assessment reflects average, long

- term base flow values , it 5 is appropriate to consider protecting against episodic drops in ANC values to a level as low as 0 6 μeq/L . Staddard et al. (2003) noted on average a 30 μeq/Ldepression of ANC between spring 7 and summer time values, indicating the need to maintain highe r base flow ANC levels to protect 8 against ANC levels below 0 μeq/L . The above considerations do not provide support for a target 9 chronic ANC level as low as 0 μeq/Lfor a standard that would protect against significant harm to 10 aquatic ecosy stems, includin g harm from episodic acidification. The Policy Assessment 11 concludes that these considerations also support a lower end of the range for consideration no 12 lower than 20 μeq/L . 13 The CASAC agreed with this conclusion in its comments on the second draft Policy 14 Assessment (Russell and Samet, 2010b). CASAC noted that “there are clear and marked 15 biological effects at ANC values near 0 μeq/L, so this is probably not an appropriate target 16 value” for the AAI. With regard to the likelihood of impairment of aquatic e cos ys tems due to 17 episodic acidification, in terms of specific target levels for chronic ANC, CASAC expressed the 18 following view: 19 Based on surface waters studied in the Northeast, decreases in ANC 20 associated with snowmelt [are] approximately 50 μeq/L . Thu s, based on these 21 studies, a long term ANC target level of 75 μeq/L would generall y guard against 22 effects from episodic acidification down to a level of about 25 μeq/L . (Russell and 23 Samet, 2010b) 24 25 26 27Draft

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161 4. Consideration of Ecosystem Response Time 1 The Policy Assessment notes that when considering a standard level to protect against 2 aquatic acidification, it is appropriate to take into account both the time period to recovery as 3 well as the potential for recovery in acid

- sensitive ecoregions. Ecos ys tems becom e adversel y 4 impacted b y acidif

yi ng deposition over long periods of time and have variable time frames and 5 abilities to recover from such perturbations. Modeling presented in the REA ( US EPA, 2009 , 6 section 4.2.4) shows the estimated ANC values for Adirondac k lakes and Shenandoah streams 7 under pre

- acidification conditions and indicates that for a small percentage of lakes and streams, 8 natural ANC levels would have been below 50 μeq/L . Therefore, for these water bodies, 9 reductions in acidifying deposition are not likely to achieve an ANC of 50 μeq/Lor greater. 10 Conversel y, for some lakes and streams the level of perturbation from long periods of acidifying 11 deposition has resulted in very low ANC values compared to estimated natural conditions. For 12 such water bodies, the time to recovery would be largel y dependent on future inputs of acidifying 13 deposition. 14 Setting a standard level in terms of a target chronic ANC level is based on the long

- term 15 response of aquatic ecosy stems. The time required for a water bod y to achieve the target ANC 16 level

-- given a decrease in ambient air concentrations of NOy and SOx and related acidifying 17 deposition such that the critical load for a target ANC is not exceeded

– is often decades if not 18 centuries. In recognition of the po tential public welfare benefits of achieving the target ANC in 19 a shorter time frame, the concept of target loads had been developed. Target loads represent the 20 depositional loading that is expected to achieve a particular level of the ecological indicator by a 21 given time. For example, to achieve an ANC level of 20 μeq/Lby 2030, it might be necessary to 22 specify

a higher target ANC level of, for example, 50 μeq/L , such that the depositional loading 23Draft

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162 would be reduced more quickly than would occur if the depo sitional loading was based on 1 achieving a target ANC level of 20 μeq/L as a long

- term equilibrium level. In this example, the 2 target ANC of 50 μeq/L would ultimately be realized many

y ears later. 3 The above considerations have implications for selecting a n appropriate standard level, in 4 that the standard level affects not only the ultimate degree of protection that would be afforded 5 by the standard, but also the time frame in which such protection would be realized. However, 6 the Policy

Assessment recogniz es that there is a great deal of heterogeneit y in response times 7 among water bodies and that there is only very limited information from dy namic modeling that 8 would help to quantify recovery time frames in areas across the country . As a consequence, 9 quant ification of a general relationship between critical loads associated with a specific long

10 term target ANC level and target loads associated with achieving the target ANC level within a 11 specific time frame is not currentl y possible. Thus, while the time f rame for recovery

is an 12 important consideration in selecting an appropriate range of levels to consider, the Policy 13 Assessment concludes that it can only be considered in a qualitative sense at this time. 14 5. Prior Examples of Target ANC Levels 15 A number o f regional organizations, states, and international organizations have 16 developed critical load frameworks to protect against acidification of sensitive aquatic 17
ecos ys tems. In considering the appropriate range of target ANC levels for consideration in this 18 review, it is informative to evaluate the target ANC levels selected b y these different 19 organizations, as well as the rationale provided in support of the selected levels. Chapter 4 of the 20 Policy

Assessment provides a detailed discussion of how critical loads have been developed and 21 used in other contexts. Specific target values and their rationales are summarized below. 22Draft

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163 The UNECE has developed critical loads in support of international emissions reduction 1 agreements. As noted in chapter 4 of the Polic y Assessment, critical loads were established to 2 protect 95 percent of surface waters in Europe from an ANC less than 20 μeq/L based on 3 protection of brown trout. Individual countries have set alternative ANC targets; for example, 4 Norway

targets an ANC of 30 μeq/Lbased on protection of Atlantic salmon. 5 Several states have established target ANC or pH values related to protection of lakes and 6 streams from acidification. While recognizing that some lakes in the Adirondacks will have a 7 naturally

low pH, the state of New York has established a target pH value of 6.5 for lakes that 8 are not naturally

below 6.5. As noted above, this level is associated with an ANC value that is 9 likely

to be between 20 and 50 μeq/L or possibl y higher. New Hampshire and Vermont have set 10 ANC targets of 60 μeq/L and 50 μeq/L , respectively . Tennessee has established site

- specific 11 target ANC values based on assessments of natural acidity , with a default value of 50 μeq/L 12 when specific data are not available. 13 Taken together, these p olicy responses to concerns about ecological effects associated 14 with aquatic acidification indicate that target ANC values between 20 and 60 μeq/Lhave been 15 selected b y states and other nations to provide protection of lakes and streams in some of the 16
more sensitive aquatic ecosy stems. 17 6. Consideration of Public Welfare Benefits 18 The point at which effects on public welfare become adverse is not defined in the CAA. 19 Characterizing a known or anticipated adverse effect to public welfare is an important 20 component of developing any secondary

NAAQS. According to the CAA, welfare effects 21 include: 22 ...effects on soils, water, crops, vegetation, manmade materials, animals, 23 wildlife, weather, visibility , and climate, damage to and deterioration of 24Draft

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164 propert y, and h azards to transportation, as well as effect on economic values and 1 on personal comfort and well

- being, whether caused by transformation, 2 conversion, or combination with other air pollutants. (CAA, section 302(h)). 3 4 While the text above lists a number of we lfare effects, the NAAQS is aimed at protection from 5 adverse effects to public welfare. Consideration of adversity to public welfare in the context of 6 the secondary

NAAQS for oxides of nitrogen and sulfur can be informed by information about 7 losses in eco sy stem services associated with acidifyingdeposition and the potential economic 8 value of those losses, as summarized above in section II. C and discussed more full y in chapter 4 9 of the Policy

Assessment. 10 Ecosy stem service losses at alternative ANC level s are difficult to enumerate. However, 11 in general there are categories of ecos ys tem services, discussed in chapter 4 of the Policy 12 Assessment, that are related to the specific ecos ys tem damages expected to occur at alternative 13 ANC levels. Losses in fish populations due to very low ANC (below 20 μeq/L ) are likel y 14 associated with significant losses in value for recreational and subsistence fishers. Many acid 15 sensitive lakes are located in areas with high levels of recreational fishing activity . For exampl e, 16 in the northeastern U.S., where nearl y 8 percent of lakes are considered acidic, more than 9 17 percent of adults participate in freshwater fishing, with an estimated value of approximately $5 18 billion in 2006. This suggests that improvements in lake fish populations are likely associated 19 with significant recreational fishing value. 20 As discussed in the Policy Assessment, inland surface waters also provide cultural 21 services such as aesthetic and existence value and educational services. To the extent that 22 piscivorous birds and other wildlife are harmed b y the absence of fish in these waters, hunting 23 and birdwatching activities are likel y to be adversely affected. A case study of the value to New 24 York residents of improving the health of lakes in the Adiron dacks found significant willingness 25Draft

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165 to pay for those improvements. When scaled to evaluate the improvement in lake health from 1 achieving ANC values of 20 to 50 μeq/L , the study implies benefits to the New York population 2 roughl y on the order of $600 mill ion per y ear (in constant 2007$). The survey administered in 3 this study

recognized that participants were thinking about the full range of services provided by 4 the lakes in question

– not just the recreational fishing services. Therefore the estimates of 5 willingness to pay

include resident’s benefits for potential hunting and birdwatching activities 6 and other ancillary

services. These results are just for New York populations. The Policy 7 Assessment concludes that if similar benefits exist for improvement s in other acid sensitive 8 lakes, the economic value to U.S. populations could be very substantial, suggesting that, at least 9 by one measure of impact on public welfare, impacts associated with ANC less than 50 μeq/L 10 may

be adverse to public welfare. 11 7. Summary of Alternative Levels 12 Based on all the above considerations, the Policy Assessment concludes that 13 consideration should be given to a range of standard levels from 20 to 75 μeq/L . The available 14 evidence indicates that target ANC levels below 20 μe q/Lwould be inadequate to protect against 15 substantial ecological effects and potential catastrophic loss of ecos ys tem function in some 16 sensitive aquatic ecosy stems. While ecological effects occur at ANC levels below 50 μeq/Lin 17 some sensitive ecos ys tems, the degree and nature of those effects are less significant than at 18 levels below 20 μeq/L. Levels at and above 50 μeq/Lwould be expected to provide additional 19 protection, although uncertainties regarding the potential for additional protection from adv erse 20 ecological effects are much larger for target ANC levels above about 75 μeq/L , as effects are 21 generall y appreciably less sensitive to changes in ANC at such higher levels. 22Draft

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166 In reaching this conclusion in the Policy Assessment, consideration was given to the 1 extent to which a target ANC level within this range would protect against episodic as well as 2 long

- term ecological effects. Levels in the mid

- to upper part of this range would be expected to 3 provide greater protection against short

- term, episodic peaks in aquatic acidification, while lower 4 levels within this range would give more weight to protection from long

- term rather than 5 episodic acidification. Similarly , levels in the mid

- to upper part of this range would be expected 6 to result in shorter time periods for recovery given the lag in ecos ys tem response in some 7 sensitive ecos ys tems relative to levels in the lower part of this range. The Policy Assessment 8 also note that this range encompasses target ANC values that have been established by vari ous 9 States and regional and international organizations to protect against acidification of aquatic 10
ecos ys tems. 11 The Policy

Assessment recognizes that the level of the standard together with the other 12 elements of the standard, including the ambient air ind icators, averaging time, and form, 13 determine the overall protectiveness of the standard. Thus, consideration of a standard level 14 should reflect the strengths and limitations of the evidence and assessments as well as the 15 inherent uncertainties in the deve lopment of each of the elements of the standard. The 16 implications of considering alternative standards, defined in terms of alternative combinations of 17 levels and percentile values that are a critical component of factor F1 in the form of the standard, 18
ar e discussed below in section III.E. Key uncertainties in the various components of the standard 19 are summarized and considered below in section III.F. 20 E. Combined Alternative Levels and Forms 21 To provide some perspective on the implications of various alt ernative standards, the 22 Policy

Assessment presented the number of acid

- sensitive ecoregions that would likely not meet 23Draft

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167 various sets of alternative standards. The alternative standards considered were based on 1 combinations of alternative target ANC levels, within the range of 20 to 75 μeq/L , and 2 alternative forms, characterized by alternative representative percentiles within the range of the 3 70 th to 90 th percentile. These alternative standards are also defined in terms of the other elements 4 of the standar d: ambient air indicators NOy and SOx, discussed above in section III.A; other 5 elements of the form of the standard, including ecoregion

- specific values for factors F1 through 6 F4 in the AAI equation, discussed above in section III.B.5; and an annual avera ging time for 7 NOy and SOx, discussed above in section III.C. With regard to the averaging time, the 8 assessment did not consider multi

- ye ar averaging of the calculated annual AAI values due to data 9 limitations, including, for example, the lack of CMAQ mode ling for multiple consecutive ye ars. 10 In this assessment, we characterize an ecoregion as likely not meeting a given alternative 11 standard if the calculated AAI value is less than the target ANC level of the standard, 12 recognizing that higher AAI values are more protective than lower values. 13 The results of this assessment are presented in Table 7

- 1a
- d in the Policy Assessment for 14 a subset of ecoregions including those characterized as acid sensitive. Calculated annual AAI 15 values at the ecoregion level are s hown for each alternative standard considered. Based on these 16 AAI values, Table 7

- 2 in the Policy Assessment summarizes the number of acid

- sensitive 17 ecoregions that would likely not meet each of the alternative standards considered. 12 Calculated 18 AAI valu es for all ecoregions categorized as relatively non - acid sensitive are shown in Table D

- 5 19 in Appendix D of the Policy Assessment. In all cases, these relatively non - acid sensitive 20 ecoregions were estimated to meet all of the alternative standards considered in this assessment. 21

12 Ta

bles 7

- 1a
- d and 7

- 2 in the Policy Assessment present assessment results for 29 ecoregions that had been initially characterized as acid sensitive. Subsequently, based on a b roader set of criteria used to characterize ecoregions as acid sensitive, as discussed above in section III.B.5.a, the set of ecoregions characterized as acid sensitive was narrow ed to include 22 ecoregions.Draft

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168 As described above, the AAI values presented in Table 7

- 1a
- d of the Policy Assessment 1 are based in part on data from 2005 CMAQ model simulations, which was used to generate 2 values for F2 through F4 in the AAI equation, as well as to estimate annual average ambient air 3 concentrations of NOy and SOx that reflect recent air quality in the absence of currentl y 4 available monitored concentrations in sensitive ecoregions across the country . Water quality and 5 hydrology

data from water bodie s within each ecoregion were also used in calculating the AAI 6 values. Such data were initially used to calculate critical loads for each water body with 7 sufficient data within an ecoregion so as to identify the nth percentile critical load representative 8 of the ecoregion used in calculating the F1 factor for the ecoregion. As expected, the number of 9 ecoregions that likel y would not meet alternative standards increases with increasing percentile 10 values and target ANC levels ( US EPA, 2011, Table 7

- 2 ). Out of 22 acid

- sensitive ecoregions, 11 the number of ecoregions that would likely not meet the alternative standards ranges from 22 for 12 the most protective alternative standard considered (75 μeq/L , 90 th percentile) to 4 for the least 13 protective alternative standard (20 μeq/L, 70 th percentile). It is apparent that both the percentile 14 and the level chosen have a strong influence, over the ranges considered, in determining the 15 number of areas that would likely not meet this set of alternative standards. 16 The Pol icy

Assessment observes that there is one grouping of these acid

- sensitive 17 ecoregions that would likely not meet almost all combinations of level and form under 18 consideration ( US EPA, 2011, Table 7

- 2 and Appendix D ). This group is made up of southern 19 Appa lachian mountain areas, including North Central Appalachians, 5.3.3; Ridge and Valley , 20 8.4.1; Central Appalachians, 8.4.2; Blue Ridge, 8.4.4; and Southwestern Appalachians, 8.4.9. In 21 addition, these ecoregions exhibit the highest amounts of exceedance re lative to alternative 22 standards. 23Draft

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169 The Northern Appalachian and Atlantic Maritime Highlands (5.3.1), which includes the 1 Adirondacks, and the Northern Lakes and Forests (5.2.1) of the upper midwest exhibit similar 2 patterns with respect to in the role of leve l and percentile in identifying regions not likely to meet 3 alternative standards, although there are considerably fewer cases compared to the regions in the 4 Appalachians. 5 In the mountainous west, the Sierra Nevada (6.2.12), Idaho Batholith (6.2.15) and th e 6 Cascades (6.2.7) ecoregions likely would not meet alternative standards in fewer cases relative to 7 eastern regions, with the Sierra Nevada ecoregion exhibiting relatively greater sensitivity 8 compared to all western regions. Onl y in the upper part of the ranges of level and percentile do 9 regions in the northern and central Rockies likel y not meet alternative standards. 10 In considering these findings, the Policy Assessment observes that the standard as 11 defined b y the AAI behaves in an intuitively logical m anner. That is, an increase in ecoregions 12 likely

not to meet the standard is associated with higher alternative levels and percentiles, both of 13 which contribute to a lower regionally representative critical load. Moreover, the areas of known 14 adverse aquatic acidification effects are identified, mostly in high elevation regions or in the 15 northern latitudes

-- the Adirondacks, Shenandoahs, northern midwest lakes and the mountainous 16 west. These results reflect the first application of a nationwide model that integrates water 17 quality

and atmospheric processes at a national scale and provides findings that are consistent 18 with our basic understanding of the extent of aquatic acidification across the U.S. What is 19 particularl y noteworthy is that this model is no t initialized with a starting ANC based on water 20 quality

data, which likely would result in a reproduction of water quality observations. Rather, 21 this standard reflects the potential of the changes in atmospheric concentrations of NOy and SOx 22 to induce lo ng - term sustained changes in surface water s ys tems. The Policy Assessment notes 23Draft

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170 that the fact that the patterns of adversity based on applying this standard are commensurate with 1 what is observed in surface water s ys tems provides confidence in the basic underlying 2 formulation of the standard. 3 The Policy

Assessment notes that the Appalachian mountain regions merit further 4 inspection as they

stand out as areas with the largest relative exceedances from a national 5 perspective. Water quality data from these regions as well as an emissions sensitivity CMAQ 6 simulation were considered to better understand the simulated behavior of these regions. The 7 maps and tables in appendix D of the Policy Assessment include paired comparisons of the 8 CMAQ 2005 and emissions sensitivity simulations. The emissions sensitivity simulation reflects 9 domain

- wide reductions in NOy and SOx emissions of 48% and 42%, respectively , relative to 10 2005 base year emissions. The Policy Assessment assumes that this emissions sensitivity 11 simul ation is indicative of future conditions. 12 The emissions sensitivity results project that man y of the regions that likel y would not 13 meet the alternative standards based on recent air quality , especiall y at alternative levels of 20 14 and 35 μeq/L , would li kely meet such standards in the future y ear scenario for the Appalachian 15 mountain regions. It is apparent that the AAI calculations are especiall y sensitive to changes in 16 SOx emissions as the Appalachian regions have the highest SOx concentrations and dep osition 17 rates ( US EPA, 2011, section 2), and the AAI equation responds as expected to modeled 18 reductions in SOx. The emissions sensitivity scenario is a prospective application of the 19 standard, in the sense that rules derived from the air quality manageme nt process result in 20 reductions of NOy

and SOx emissions. Expected emission changes over the next two decades 21 should be far greater than the 42 and 48% SOx and NOy reductions used in this analy sis, with a 22 consequent further reduction in areas that would likely not meet alternative standards. 23Draft

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171 The Appalachian mountain regions generally have low DOC levels, average runoff rates, 1 moderatel y low base cation supply and highl y elevated sulfate concentrations. Collectively, 2 those attributes do not suggest natur all y acidic conditions as the availability of anthropogenic 3 contributions of mineral acids is likely responsible for observed low ANC values in those 4 regions. 5 The Policy

Assessment notes the Sierra Nevada region as an interesting case study , as it 6 has som e of the lowest critical load values nationally ( US EPA, 2011, Table D

- 3 ). Water quality 7 data indicate extremel y l ow sulfate, as expected given the relativel y low SO 2 emissions in the 8 western U.S. Extremely low base cation suppl y and low Neco, which mitigate the effect of 9 nitrogen deposition, explain the low critical load values. Low Neco values appear to associate 10 well with high elevation western U.S. regions, perh aps reflecting the more arid and reduced 11 vegetation density

relative to eastern U.S. regions. The proximity to high level nitrogen 12 emissions combined with very low base cation supply

explains the cases where the Sierra region 13 likely

does not meet alternati ve standards. Because Neco values are low in the Sierras, the 14 sy stem responds effectively to reductions of NOx emissions, as illustrated in the maps and tables 15 of Appendix D of the Policy Assessment. Although Neco affords protection from the acidifying 16
e ffects of nitrogen deposition, the availability of excessive nitrogen neutralization capacity also 17 means that reductions in nitrogen are not as effective as reductions in SOx in reducing the 18 calculated AAI. 19 In reviewing these results, the Policy Assessment observes that the anal ys is of the 20 alternative combinations of level and form presented provide context for considering the impact 21 of different standards. Since the AAI equation has been newl y developed in the Policy 22 Assessment, these examples of estimated exceedances help to address the question of whether 23Draft

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172 the AAI equation responds in a reasonable manner with regard to identifying areas of concern 1 and to prospective changes in atmospheric conditions likely to result from future emissions 2 reduction strat egies. The Policy Assessment concludes that the behavior of the AAI calculations 3 is both reasonable and explainable, which the Policy Assessment concludes serves to increase 4 confidence in considering a standard defined in terms of the AAI. 5 F. Characteriza tion of Uncertainties 6 This section summarizes discussions of the results of anal ys es and assessments, presented 7 more fully

in the Policy

Assessment ( US EPA, 2011 , section 7.6 and Appendices F and G), 8 intended to address the relative confidence associated with the linked atmospheric

- ecological 9 effects sy stem described above. An overview of uncertainties is presented in the context of the 10 major structural components underl yi ng the standard, as well as with regard to areas of relativel y 11 high uncertaint y. Th e section closes with a discussion of data gaps and uncertainties associated 12 with the use of ecological and atmospheric modeling to specify the factors in the AAI equation, 13 which can be used to guide future field programs and longer

- term research efforts. 14 1. Overview of Uncertainty 15 As discussed in the Policy Assessment ( US EPA, 2011, Table 7
- 3), there is relativel y low 16 uncertaint y with regard to the conceptual formulation of the overall structure of the AAI

- based 17 standard that incorporat es the major associations linking biological effects to air concentrations. 18 Based on the strength of the evidence that links species richness and mortality to water quality , 19 the associations are strongl y causal and without any obvious confounding influence. The strong 20 association between the ecosy stem indicator (ANC) and the causative water chemistry species 21 (dissolved aluminum and hydrogen ion) reinforces the confidence in the linkage between 22 deposition of nitrogen and sulfur and effects . This strong asso ciation between ANC and effects 23Draft

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173 is supported by

a sound mechanistic foundation between deposition and ANC. The same 1 mechanistic strength holds true for the relationship between ambient air levels of nitrogen and 2 sulfur and deposition, which completes the linkage from ambient air indicators through 3 deposition to ecological effects. 4 There are relativel y higher uncertainties, however, in considering specific elements 5 within the structure of an AAI
- based standard, including the deposition of SOx, NOy , an d NHx 6 as well as the critical load

- related component, each of which can vary within and across 7 ecoregions. Overall sy stem uncertaint y relates not just to the uncertainty in each such element, 8 but also to the combined uncertainties that result from linking these elements together within the 9 AAI
- based structure. Some of these elements

– including , for example, dry deposition, pre

10 industrial base cation production , and reduced nitrogen deposition

– are estimated with less 11 confidence than other elements ( US EPA, 2011, Table 7.3 ). The uncertainties associated with all 12 of these elements, and the combination of these elements through the AAI equation, are 13 discussed below and in the following sections related to measured data gaps and modeled 14 processes for both air quality and water quality

. 15 The lack of observed dry deposition data is constrained b y resources and the lack of 16 efficient measurement technologies. Progress in reducing uncertainties in dry deposition will 17 depend on improved atmospheric concentration data and direct deposition flux measurements of 18 the relevant suite of NOy and SOx species

. 19 Pre
- industrial base cation productivity by definition is not observable. Contemporary 20 observations and inter

- model comparisons are useful tools that would help re duce the uncertainty 21 in estimates of preindustrial base cation productivity used in the AAI equation

. In characterizing 22 contemporary

base cation flux using basic water quality measurements (i.e., major anioni and 23Draft

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174 cation species as defined in equation 2.11 in the Policy Assessment) , it is reasonable to assume 1 that a major component of contemporary base cation flux is associated with pre

- industrial 2 weathering rates. To the extent that multiple models converge on similar solutions, greater 3 confidence in esti mating pre

- industrial base cation production would be achieved. 4 Characterization of reduced nitrogen (NHx) deposition has been evolving over the last 5 decade. The relativel y high uncertaint y in characterizing NHx deposition is due to both the lack 6 of fiel d measurements and the inherent complexity of characterizing NHx with respect to source 7 emissions and dry

deposition. Because ammonia emissions are generated through a combination 8 of man

- made and biological activities, and ammonia is semi

- volatile, the ab ility to characterize 9 spatial and temporal distributions of NHx concentration s and deposition patterns is challenging. 10 While direct measurement of NHx deposition is resource intensive because of the diffuse nature 11 of sources (i.e., area

- wide and not point sources), there have been more frequent deposition flux 12 studies, relative to other nitrogen species, that enable both the estimation of emissions and dry 13 deposition. Also, while ammonia has a relativel y high deposition velocit y and traditionally was 14 thou ght to deposit close to the emissions release areas, the semi

- volatile nature of ammonia 15 results in re

- entrainment back into the lower boundary lay er resulting in a more dispersed 16 concentration pattern exhibiting transport ty pe characteristics similar to l onger lived atmospheric 17 species. These inherent complexities in source characterization and ambient concentration 18 patterns raise the uncertainty level of NHx in general. However, the Policy Assessment notes 19 that progress is being made in measuring ammoni a with cost efficient samplers and anticipates 20 the gradual evolution of a spatially robust ammonia sampling network that would help support 21 anal ys es to reduce underlyinguncertainties in NHx deposition. Also, from an aquatic 22 acidification perspective, NHx is not as important a driver relative to NOy and SOx in the 23Draft

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175 mountainous areas in the eastern U
. S. However, the relative importance of NHx is likely to 1 increase over time, in light of air quality rules in place designed to reduce emissions of NOy and 2 SOx. 3 2. Uncertainties Associated with Data Gaps 4 In summarizing uncertainties with respect to available measurement data and the use of 5 ecological and atmospheric models, the Policy Assessment indentified data gaps and model 6 uncertainties in relative terms by comparing, for example, the relative richness of data between 7 geographic areas or environmental media. With regard to relevant air quality measurements, the 8 Policy

Assessment notes that such measurements are relativel y sparse in the western U.S. While 9 th e spatial extent of CASTNET coverage has gradually incorporated western U.S. locations with 10 support from the National Parks Service (NPS), the relative density of monitoring sites is much 11 less than that in the eastern U.S. This relative disparity in spati al density

of monitors is 12 exacerbated as air quality patterns in the mountainous west generally exhibit greater spatial 13 heterogeneity

due to dramatic elevation gradients that impact meteorology and air mass flow 14 patterns. Similarly , water quality data cov erage is far more comprehensive in the eastern U.S. 15 relative to the west 16 Measurements of NOy notably are lacking in both eastern and western acid

- sensitive 17 ecoregions. This adds uncertaint y to the use of the AAI equation as the lack of NOy data limits 18 efforts to evaluate air quality model ing of NOy that is the basis for quantifying factor F3 in the 19 AAI equation . The lack of NOy measurements also limits efforts to characterize the variability 20 and representative ness of modeled NOy concentrations within and across ecoregions

. Currently , 21 the Agency ’s ability to define the protection likely to be afforded by

alternative standards (in 22 terms of alternative levels and percentiles) is compromised by the lack of a full set of ambient air 23Draft

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176 quality indicator measur ements , notably including NOy , throughout sensitive ecoregions across 1 the U.S

. 2 Further, obtaining measurement of the dominant species that comprise NOy ( HNO 3 , true 3 NO 2 , NO , p
–

NO 3 , and PAN) would be useful to evaluate performance of NOy samplers. 4 Be yon d the more well known dominant components of NOy , research efforts would be needed to 5 characterize total reactive nitrogen that may include significant amounts of organicall y - b ound 6 nitrogen (beyond PAN) which is poorly understood with regard to emission so urces and 7 concentration levels. 8 Field measurements of NHx have been extremel y limited, but have begun to be enhanced 9 through the National Atmospheric Deposition Program’s (NADP) passive ammonia network 10 (AMoN). AMoN measures ammonia at over 50 sites, with more than 35 at CASTNET locations. 11 Enhanced spatial coverage of reduced nitrogen measurements , particularl y t o understand within 12 and across ecoregion variability , and the inclusion of some continuous observations would 13 provide a better understanding of th e uncertainty in the F2 factor in the AAI equation and of the s 14 representativeness of modeled NHx deposition within and across ecoregions. 15 With regard to water quality data, the Policy

Assessment notes that such data are 16 ty picall y limited relative to air quality data sets, and are also relativel y sparse in the western U.S. 17 The TIME/LTM water quality sampling program in the eastern U.S. (as described in chapter 2 18 of the Policy

Assessment) is an appropriate complement to national air monitoring programs as it 19 affords consistency across water bodies in terms of sampling frequency and anal ys is protocols. 20 Consideration should be given to extending the TIME/L TM design to all acid sensitive 21 ecoregions, with priorit y for areas in the western mountains that are d ata limited and showing 22 initial signs of adversit y particularl y with respect to aquatic acidification. The lack of a 23Draft

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177 regulatory

requirement for TIME/LTM often jeopardizes funding support of this resource that is 1 especiall y valuable and cost effective. Wh ile there are several state and local agency water 2 quality

data bases, it is unclear the extent to which differences in sampling, chemical analy sis 3 and reporting protocols would impact the use of such data for the purpose of better 4 understanding the degree of protectiveness that would be afforded b y an AAI
- based standard 5 within sensitive ecoregions across the country . In addition, our understanding of water quality in 6 Alaska and Hawaii and the acid sensitiv ity of their ecoregions is particularly limited

. 7 Water quality data and modeling support the standard setting process. As more water 8 bodies are sampled, the critical load data bases would expand, enabling clearer delineation of 9 ecoregion representative critical loads in terms of the nth percentile. T his w ould provide more 10 refined characterization of the degree of protection afforded by a given standard. Longer term, 11 the availability

of water quality

trend data (annual to monthly sampled) would support 12 accountability

assessments that examine if an eco region’s response to air management efforts is 13 as predicted b y earlier model forecasting. The most obvious example is the long

- term response 14 of water quality

ANC change to changes in calculated AAI, deposition, ambient NOy and SOx 15 concentrations, and emis sions. In addition, water quality trends data provide a basis for 16 evaluating and improving the parameterizations of processes in critical load models applied at 17 the ecoregion scale related to nitrogen retention and base cation suppl y. A better understand ing 18 of soil processes , especially in the southern Appalachians, would enhance efforts to examin e the 19 variability

within ecoregions of the soil

- based adsorption and exchange processes which 20 moderate the suppl y of major cations and anions to surface waters and strongl y influence the 21 response of surface water ANC to changes in deposition of nitrogen and sulfur. 22Draft

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178 3. Uncertainties in Modeled Processes 1 As discussed in the Policy Assessment, from an uncertaint y perspective, gaps in field 2 measurement data are relat ed to uncertainties in modeled processes and in the specific 3 application of such models

. As noted above, processes that are embodied in an AAI

- based 4 standard are modeled using the CMAQ atmospheric model and steady state ecological models. 5 These models ar e characterized in the Integrated Science Assessment as being well established 6 and they

have undergone extensive peer review. Nonetheless, the application of these models for 7 purposes of specifying the factors in the AAI equation, on an ecoregion scale, i s a new 8 application that introduces uncertainties, as noted below, especially in areas with limited 9 observational data that can be used to evaluate this specific application. Understanding 10 uncertainties in relevant modeled process thus involves considerat ion of the uncertainties 11 associated with appl yi ng each model as well as the combination of these uncertainties as the 12 models are applied in combination within the AAI framework. 13 With regard to the application of CMAQ for purposes of use in an AAI

- based standard, 14 the modeling of dry deposition has been identified as having a relativel y high degree of 15 uncertaint y. Due to a combination of sy stem complexity and resource constraints, there is no 16 routine observational basis for directl y comparing modeled dry dep osition and measurements. 17 Periodic dry

deposition flux experiments covering a variet y of vegetation, surfaces and 18 meteorology

across seasons would enable a more robust evaluation of modeled deposition of 19 nitrogen and sulfur. Given the difficult y in acqui ring dry deposition observations, it becomes 20 especiall y important to evaluate the model’s ability to capture temporal and spatial ambient air 21 patterns of individual nitrogen and sulfur species which are used to drive dry deposition 22 calculations in models. For example, reducing a generall y acknowledged positive bias in model

- 23Draft

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179 predicted SO 2 relative to observations is especiall y relevant to the AAI

- based standard, as SO 2 1 deposition is a dominant contributor to total acidifyingdeposition in the eastern U.S. With 2 respect to oxidized nitrogen, observations of individual NOy species are important as air qualit y 3 models calculate the individual deposition of each species. The modeled transference ratios, 4 T NOy and T SOx used in factors F3 and F4 rel y on CMAQ’s abil ity to characterize both deposition 5 and concentration. Consequently , a better understanding of the variability of these factors 6 within and across ecoregions could be achieved b y improved availability of measured ambient 7 concentration s and deposition observations. 8 Steady

state biogeochemical ecos ys tem modeling is used to develop critical load 9 estimates that are incorporated in the AAI equation through factor F1. Consequently , the Policy 10 Assessment notes that an estimate of the tempora l response of surface water ANC to deposition 11 and air concentration changes is not directly available. Lacking a predicted temporal response 12 impairs the ability

to conduct accountability assessments down to the effects level. 13 Accountability

assessments w ould examine the response of each step in the emissions source 14 through air concentration

– deposition

-- surface water qualit y
– biota continuum. The steady 15 state assumption at the ecosy stem level does not impair accountability assessments through the 16 air concentration/deposition range of that continuum. However, i n using steady state ecos ys tem 17 modeling, several assumptions are made relative to the long

- term importance of processes 18 related to soil adsorption of major ions and ecos ys tem nitrogen d yn amics. Because these models 19 often were developed and applied in glaciated areas with relativel y thin and organicall y rich 20 soils, their applicability is relativel y more uncertain in areas such as those in the non

- glaciated 21 clay

- based soil regions of the central Ap palachians. Consequently , it is desirable to develop the 22 information bases to drive simple dy namic ecos ys tem models that incorporate more detailed 23Draft

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180 treatment of subsurface processes, such as adsorption and exchange processes and sulfate 1 absorption. 2 4. Appl

ying Knowledge of Uncertainties 3 An understanding of the relative uncertainties in a sy stem assists in setting priorities for 4 data collection efforts and research, with the expectation that such efforts would reduce 5 uncertainties over time and afford great er confidence in applications of an AAI

- based standard. 6 Because of the uniquel y wide breadth of pollutants and environmental media addressed b y an 7 AAI
- based multi

- pollutant standard, there are a wide range of uncertainties that are important to 8 consider relative to single pollutant standards that ty picall y address onl y direct effects of ambient 9 air exposures. For an AAI
- b ased standard, an improved understanding of the uncertainties 10 across the various modeled processes at the ecoregion scale would lead t o greater confidence in 11 the degree of protection afforded b y the standard. 12 The Policy

Assessment notes that there is general ly low uncertaint y with regard to the 13 conceptual development and related major components of this standard. In recognizing the 14 scientific soundness of the basic structure of this standard, the Policy Assessment notes that 15 future efforts would be appropriatel y directed at expanding the availability of relevant data for 16 ecoregion

- specific e valuati on and application of the relevant m odeling of ecological and 17 atmospheric processes, as identified above . Such efforts would further support consideration of 18 an AAI
- based standard and would guide field studies and anal ys es designed to improve the 19 longer

- term confidence in such a standard. 20 G

. CASAC Advice 21 The CASAC has advised EPA concerning the Integrated Science Assessment, the Risk 22 and Exposure Assessment, and the Policy Assessment. CASAC has endorsed EPA’s 23Draft

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181 interpretation of the science embodied in the Integrated Science Assessment and the assessment 1 approaches and conclusions incorporated in the Risk and Exposure Assessment

. 2 Most recentl y, CASAC has considered the information in the final Policy Assessment in 3 providing its recommendations on the review of the new multi

- pollutant standar d developed in 4 that document and discussed above (Russell and Samet, 2011 a). In so doing, CASAC has 5 expressed general support for the conceptual framework of the standard based on the underl yi ng 6 scientific information, as well as for the conclusions in th e Policy Assessment with regard to 7 indicators, form, averaging time, and level of the standard that are appropriate for consideration 8 by the Agency

in reaching decisions on the review of the secondary NAAQS for oxides of 9 nitrogen and sulfur: 10 The final Poli cy Assessment clearl y s ets out the basis for the recommended ranges 11 for each of the four elements (indicator, averaging time, level and form) of a 12 potential NAAQS that uses ambient air indicators to address the combined effects 13 of oxides of nitrogen and oxides of sulfur on aquatic ecos ys tems, primaril y 14 streams and lakes. As requested in our previous letters, the Policy Assessment also 15 describes the implications of choosing specific combinations of elements and 16 provides numerous maps and tabular estimates of the spatial e xtent and degree of 17 severit y of NAAQS exceedances expected to result from possible combinations of 18 the elements of the standard. 19 20 We believe this final PA is appropriate for use in determining a secondary 21 standard to help protect aquatic ecos ys tems from acidifying deposition of oxides 22 of sulfur and nitrogen. EPA staff has done a commendable job developing the 23 innovative Aquatic Acidi fication Index (AAI), which provides a framework for a 24 national standard based on ambient concentrations that also takes into account 25 regional differences in sensitivities of ecosy stems across the country to effects of 26 acidifying deposition. (Russell and S amet, 2011 a) 27 28 CASAC also recommended that as EPA moves forward in the regulatory process “some 29 attention should be given to our residual concern that the available data may reflect the more 30 sensitive water bodies and thus, the selection of percentiles of waterbodies to be protected could 31Draft

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182 be conservativel y biased” (Russell and Samet, 2011 a). In additional, CASAC found some 1 improvements could be made to the uncertainty analy sis, as noted below. 2 With respect to indicators, CASAC supports the use of SOx and N Oy as ambient air indicators 3 (discussed above in section III.A) and ANC as the ecological indicator (discussed above in 4 section III.B.1): 5 The use of total reactive oxidized nitrogen (NOy) and sulfur oxides (SOx) as atmospheric 6 indicators of oxidized nitr ogen (N) and sulfur (S) atmospheric concentrations is well justified. 7 The use in the Aquatic Acidification Index (AAI) of NOy and SOx as atmospheric indicators of 8 N and S concentrations is useful and corresponds with other efforts by EPA. As we have state d 9 previously , CASAC also agrees that acid neutralizing capacit y (ANC) is the most appropriate 10 ecological indicator of aquatic ecos ys tem response and resiliency to acidification. (Russell and 11 Samet, 2011 a) 12 With respect to the form of the standard (discusse d above in section III.B), CASAC 13 stated the following: 14 EPA has developed the AAI, an innovative “form” of the NAAQS itself that 15 incorporates the multi

- pollutant, multi

- media, environmentally modified, 16 geographically variable nature of SOx/NOy deposition

- re lated aquatic 17 acidification effects. With the caveats noted below, CASAC believes that this 18 form of the NAAQS as described in the final Policy Assessment is consistent with 19 and directl y reflective of current scientific understanding of effects of acidifyin g 20 deposition on aquatic ecosy stems. (Russell and Samet, 2011 a) 21 22 CASAC agrees that the spatial components of the form in the Policy Assessment 23 are reasonable and that use of Omernick’s ecoregions (Level III) is appropriate for 24 a secondary NAAQs intended to protect the aquatic environment from 25 acidification . . . (Russell and Samet, 2011 a) 26 27 The “caveats” noted b y CASAC include a recognition of the importance of continuing to 28 evaluate the performance of the CMAQ and ecological models to account for model 29Draft

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183 unc ertainties and to make the model

- dependent factors in the AAI more transparent. In addition, 1 CASAC noted that the role of DOC and its effects on ANC would benefit from further 2 refinement and clarification (Russell and Samet, 2011 a). While CASAC expressed the view that 3 the “division of ecoregions into ‘sensitive’ and ‘non

- sensitive’ subsets, with a more protective 4 percentile applied to the sensitive areas, also seems reasonable” (Russell and Samet, 2011 a), 5 CASAC also noted that there was the need for great er clarity in specifying how appropriate 6 screening criteria would be applied in assigning ecoregions to these categories. Further, CASAC 7 identified potential biases in critical load calculations and in the regional representativeness of 8 available water che mistry data, leading to the observation that a given percentile of the 9 distribution of estimated critical loads may be protective of a higher percentage of surface waters 10 in some regions (Russell and Samet, 2011 a). 11 With respect to averaging time (discuss ed above in section III.C), CASAC stated the 12 following: 13 Considering the cumulative nature of the long

- term adverse ecological effects and 14 the y ear - to - year variability of atmospheric conditions (mainly in the amount of 15 precipitation), CASAC concurs with EPA that an averaging time of three to five 16 ye ars for the AAI parameters is appropriate. A longer averaging time would mask 17 possible trends of AAI, while a shorter averaging time would make the AAI being 18 more influenced b y the conditions of the particular yea rs selected. (Russell and 19 Samet, 2011 a) 20 21 With respect to level as well as the combination of level and form as they are presented 22 as alternative standards (discussed above in sections III.D

- E), CASAC stated the following: 23 CASAC agrees with EPA staff’s rec ommendation that the “level” of the 24 alternative AAI standards should be within the range of 20 and 75 μeq/L. We also 25 recognize that both the “level” and the form of any AAIstandard are so closely 26 linked in their effectiveness that these two elements shoul d be considered 27 together. (Russell and Samet, 2011 a) 28 29Draft

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184 When considered in isolation, it is difficult to evaluate the logic or implications 1 of selecting from percentiles (70th to 90th) of the distribution of estimated critical 2 loads for lakes in sensitive e coregions to determine an acceptable amount of 3 deposition for a given ecoregion. However, when these percentile ranges are 4 combined with alternative levels within the staff

- recommended ANC range of 20 5 to 75 microequivalents per liter (μeq/L), the results u sing the AAI point to the 6 ecoregions across the country that would be expected to require additional 7 protection from acidifying deposition. Reasonable choices were made in 8 developing the form. The number of acid sensitive regions not likely to meet the 9 s tandard will be affected both by choice of ANC level and the percentile of the 10 distribution of critical loads for lakes to meet alternative ANC levels in each 11 region. These combined recommendations provide the Administrator with a 12 broad but reasonable ran ge of minimally to substantially

protective options for the 13 standard. (Russell and Samet, 2011 a) 14 15 CASAC also commented on EPA’s uncertaint y analysis, and provided advice on areas 16 requiring further clarification in the proposed rule and future research. C ASAC found it 17 “difficult to judge the adequacy of the uncertaint y anal ys is performed by EPA because of lack of 18 details on data inputs and the methodology used, and lack of clarity

in presentation” (Russell and 19 Samet, 2011 a). In particular, CASAC identifie d the need for more thorough model evaluations 20 of critical load and atmospheric modeling, recognizing the important role of models as they are 21 incorporated in the form of the standard. In light of the innovative nature of the standard 22 developed in the Pol icy Assessment, CASAC identified “a number of areas that should be the 23 focus of further research” (Russell and Samet, 2011 a). While CASAC recognized that EPA staff 24 was able to address some of the issues in the Policy Assessment, they

also noted areas “that 25 would benefit from further study or consideration in potential revisions or modifications to the 26 form of the standard.” Such research areas include “sulfur retention and mobilization in the 27 soils, aluminum availability , soil versus water acidification and ecos ys tem recovery times.” 28 Further, CASAC encouraged future efforts to monitor individual ambient nitrogen species, 29Draft

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185 which would help inform further CMAQ evaluations and the specification of model

- derived 1 elements in the AAI equation (Russell and Samet, 2011 a). 2 H. Administrator’s Proposed Conclusions 3 Having concluded that the existing NO 2 and SO 2 secondary standards are neither 4 sufficiently

protective nor appropriate to address deposition

- related effects associated with 5 oxides of nitrogen and sulfur (se ction II.D above), the Administrator has considered whether it is 6 appropriate at this time to set a new multi

- pollutant standard for that purpose. In particular, she 7 has focused on the new standard developed in the Policy Assessment and reviewed by CASAC, 8 as discussed above. In so doing, the Administrator first considered the extent to which there is a 9 scientific basis for development of such a standard, specificall y with regard to a standard that 10 would provide protection from deposition

- related aquatic a cidification in sensitive aquatic 11
ecos ys tems in areas across the country . As discussed above, the Administrator notes that the 12 Integrated Science Assessment concludes that the available scientific evidence is sufficient to 13 infer a causal relationship betw een acidifying deposition of nitrogen and sulfur in aquatic 14
ecos ys tems, and that the deposition of oxides of nitrogen and sulfur both cause such acidification 15 under current conditions in the U.S. Further, the Integrated Science Assessment concludes that 16 there are well

- established water qualit y and biological indicators of aquatic acidification as well 17 as well

- established models that link deposition, water qualit y, and effects on ecos ys tem biota, 18 and that ecos ys tem sensitivity to acidification varies acros s the country according to present and 19 historic nitrogen and sulfur deposition as well as geologic, soil, vegetative, and hydrologic 20 factors. Based on these considerations, the Administrator agrees with the conclusion in the 21 Policy

Assessment, and support ed b y CASAC, that there is a strong scientific basis for 22 development of such a standard. 23Draft

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186 The Administrator also recognizes that the conceptual framework for an ecologically 1 relevant, multi

- pollutant standard, which was initially explored in the Risk and E xposure 2 Assessment and further developed in the Policy Assessment, builds on the information in the 3 Integrated Science Assessment. She notes that the structure of the standard addresses the 4 combined effects of deposition from oxides of nitrogen and sulfur by characterizing the linkages 5 between ambient concentrations, deposition, and aquatic acidification, and that the structure of 6 the standard takes into account relevant variations in these linkages across the country . Further, 7 she recognizes the standard depends on atmospheric and ecological modeling to specify the 8 terms of an equation that incorporates these linkages. This equation, which defines an aquatic 9 acidification index (AAI), has the effect of translating spatiall y variable ecological effects into a 10 national standard. She recognizes that while the standard is innovative and unique, the structure 11 of the standard is well grounded in the science underlying the relationships between ambient 12 concentrations of oxides of nitrogen and sulfur and the aqu atic acidification related to deposition 13 of nitrogen and sulfur associated with such ambient concentrations. 14 While the Administrator recognizes the strong scientific foundation for the structure of 15 the AAI standard, she also recognizes a number of inherent uncertainties and complexities that 16 are relevant to the question of whether it is appropriate at this time to set such a standard. As 17 discussed above, these uncertainties and complexities generall y relate not to the structure of the 18 standard, but to the quantification of the various elements of the standard, such as the F factors 19 discussed earlier in this section and their representativeness at an ecoregion scale

. These 20 uncertainties and complexities currentl y limit efforts to characterize the de gree of protectiveness 21 that would be afforded by such a standard, within the ranges of levels and forms identified in the 22 Policy

Assessment, and the representativeness of F factors in the AAI equation described above 23Draft

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187 and in the Policy

Assessment. These important uncertainties have been generall y categorized as 1 limitations in available field data as well as uncertainties that are related to reliance on the 2 application of ecological and atmospheric modeling at the ecoregion scale to specify the various 3 eleme nts of the AAI. 4 With regard to data limitations, the Administrator observes that there are several 5 important limitations in the available data upon which elements of the AAI are based. For 6 example, while ambient measurements of NOy are made as part of a national monitoring 7 network, the monitors are not located in locations that are representative of sensitive aquatic 8
ecos ys tems. While air and water qualit y data are generall y available in areas in the eastern U.S., 9 there is relativel y sparse coverage in m ountainous western areas where a number of sensitive 10 aquatic ecosy stems are located. Further, even in areas where relevant data are available, small 11 sample sizes impede efforts to characterize the representativeness of the available data, which 12 was noted by

CASAC as being of particular concern. Also, measurements of reduced forms of 13 nitrogen are available from only a small number of monitoring sites, and emission inventories for 14 reduced forms of nitrogen used in atmospheric modeling are subject to conside rable uncertaint y. 15 With regard to uncertainties related to the use of ecological and atmospheric modeling, 16 the Administrator notes in particular that model results are difficult to evaluate due to a lack of 17 relevant observational data. For example, relat ively large uncertainties are introduced by a lack 18 of data with regard to pre

- industrial environmental conditions and other parameters that are 19 necessary

inputs to critical load models that are the basis for factor F1 in the AAI equation. 20 Also, observatio nal data are not generall y available to evaluate the modeled relationships 21 between nitrogen and sulfur in the ambient air and associated deposition, which are the basis for 22 the other factors (i.e., F2, F3, and F4) in the AAI equation. 23Draft

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188 In combination, thes e limitations and uncertainties result in a considerable degree of 1 uncertaint y as to how well the quantified elements of the AAI standard would predict the actual 2 relationship between varying ambient concentrations of oxides of nitrogen and sulfur and stea dy 3 state ANC levels across the distribution of water bodies within the various ecoregions in the U.S. 4 Because of this, there is considerable uncertaint y as to the actual degree of protectiveness that 5 such a standard would provide, especiall y for acid

- sens itive ecoregions. The Administrator 6 recognizes that the AAI equation, with factors quantified as discussed above and described more 7 fully in the Policy

Assessment, generall y performs well in identifying areas of the country that 8 are sensitive to such acid ifying deposition and indicates, as expected, that lower ambient levels 9 of oxides of nitrogen and sulfur would lead to higher calculated AAI values. However, the 10 uncertainties discussed here concern the actual degree of protection that would be afforded s uch 11 areas b y an y specific target ANC level and percentile of water bodies that would be chosen in 12 setting a new AAI standard. 13 In considering these uncertainties, the Administrator notes that CASAC acknowledged 14 that important uncertainties remain that woul d benefit from further study and data collection 15 efforts, which might lead to potential revisions or modifications to the form of the standard 16 developed in the Policy Assessment. She also notes that CASAC encouraged the Agency to 17 engage in future monitori ng and model evaluation efforts to help inform the specification of 18 model

- derived elements in the AAI equation. 19 Based on the above considerations, the Administrator has strong reservations about the 20 appropriateness of setting a new multi

- pollutant standar d to address deposition

- related effects of 21 oxides of nitrogen and sulfur on aquatic acidification at this time. Setting a NAAQS generall y 22 involves consideration of the degree of uncertainties in the science and other information, such 23Draft

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189 as gaps in the relev ant data and, in this case, limitations in the evaluation of the application of 1 relevant ecological and atmospheric models at an ecoregion scale

. The issue here is the 2 appropriate public welfare policy judgment to be made in light of the specific circumst ances in 3 this review. As noted above, it is not a question of uncertainties about the scientific soundness of 4 the structure of the AAI, but instead uncertainties in the quantification and representativeness of 5 the elements of the AAI as they vary in ecore gions across the country . At present, these 6 uncertainties interfere with understanding the degree of protectiveness that would be afforded 7 various ecoregions across the country by a new standard defined in terms of a specific 8 nationwide target ANC level and a specific percentile of water bodies for acid

- sensitive 9 ecoregions. The Administrator has considered whether these uncertainties could be 10 appropriatel y accounted for by choosing either a more or less protective target ANC level and 11 percentile of water bodies than would otherwise be chosen if the uncertainties did not 12 substantially

limit the confidence that can appropriately be ascribed to the quantification of the 13 AAI elements. However, in the Administrator’s judgment, the uncertainties are of such nature 14 and magnitude that there is no reasoned way to choose such a specific nationwide target ANC 15 level or percentile of water bodies that would appropriately account for the uncertainties, since 16 neither the direction nor the magnitude of change from the ta rget level and percentile that would 17 otherwise be chosen can reasonabl y be ascertained at this time. 18 Based on the above considerations, the Administrator judges that the current limitations 19 in relevant data and the uncertainties associated with specifyin g the elements of the AAI based 20 on modeled factors are of such nature and degree as to prevent her from reaching a reasoned 21 decision such that she is adequatel y confident as to what level and form (in terms of a selected 22 percentile) of such a standard woul d provide an intended degree of protection of public welfare. 23Draft

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190 While acknowledging that CASAC supported moving forward to establish the standard 1 developed in the Policy Assessment, the Administrator also observes that CASAC supported 2 conducting further fi eld studies that would better inform the continued development or 3 modification of such as standard. Given the large uncertainties and complexities inherent in 4 quantifying the elements of such a standard, largely deriving from the unprecedented nature of 5 the standard under consideration in this review, and having fully considered CASAC’s advice, 6 the Administrator provisionally concludes that it is premature to set a new, multi

- pollutant 7 secondary

standard for oxides of nitrogen and sulfur at this time, and as such she is not proposing 8 such a new secondary standard. 9 Instead, the Administrator determines that the Agency should undertake a field pilot 10 program to gather additional data, and that it is appropriate that such a program be undertaken 11 before, rather than after, reaching a decision to set such a standard. As described below in 12 section IV, the purpose of the program is to collect and anal yz e data so as to enhance our 13 understanding of the degree of protectiveness that would likely be afforded by

a stan dard based 14 on the AAI as developed in the Policy Assessment. Data generated b y this field program will 15 also support development of an appropriate monitoring network that would work in concert with 16 such a standard to result in the intended degree of protec tion. The data and anal ys es generated as 17 a result of this program will serve to inform the next review of the NAAQS for oxides of 18 nitrogen and sulfur. 19 Based on the above considerations , the Administrator is deferring the setting of a new 20 multi

- pollutant secondary standard in this review, and is proposing not to set such a secondary 21 standard at this time. In reaching this decision, the Administrator recognizes that the new NO 2 22 and SO 2 primary

1
- hour standards set in 2010, while not ecologically relevant, will nonetheless 23Draft

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191 result in reductions in oxides of nitrogen and sulfur that will directionally benefit the 1 environment by

reducing NOy

and SOx deposition to sensitive ecosy stems. The Administrator 2 observes that an y such environmental benefits will occur a s a result of strategies designed to 3 meet the new primary standards, and that in this case no additional benefit would result from 4 setting new secondary standards identical to those new primary standards. 5 Having reached this decision, the Administrator al so recognizes that a new, innovative 6 AAI
- based standard would raise significant implementation issues that would need to be 7 addressed consistent with the CAA requirements for implementation

- related actions following 8 the setting of a new NAAQS. It will tak e time to address these issues, during which the Agency 9 will be conducting a field pilot program to gather relevant data and the environment will benefit 10 from reductions in oxides of nitrogen and sulfur resulting from the new NO 2 and SO 2 primary 11 standards, as noted above , as well as reductions expected to be achieved from EPA’s proposed 12 Transport Rule and Mercury and Air Toxics standards. These implementation

- related issues are 13 discussed in more detail below in section IV.A.5. 14 The Administrator solicits comment on all aspects of this proposed decision, including 15 the framework and elements of a multi

- pollutant standard for oxides of nitrogen and sulfur to 16 address deposition

- related effects on sensitive ecosy stems, with a focus on aquatic acidification, 17 and the uncertainties and complexities associated with the development of such a standard at this 18 time. The Administrator also solicits comment on the field pilot program and related monitoring 19 methods as discussed below in section IV. 20 IV. Field Pilot Progra m and Ambient Monitoring 21 This section describes EPA’s plans for a field pilot program and the evaluation of 22 monitoring methods for ambient air indicators of oxides of nitrogen (NOy) and oxides of sulfur 23Draft

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192 (SOx) to implement the Administrator’s decision to undertake such a field monitoring program 1 in conjunction with her decision to propose not to set a new multi

- pollutant secondary standard 2 in this review, as discussed above in section III.H. As noted above and discussed below in 3 section IV.A, the field pil ot program is intended to collect and analyze data so as to enhance our 4 understanding of the degree of protectiveness that would likely be afforded by a standard based on 5 the AAI as developed in the Policy Assessment. Data generated by this field program would also 6 support development of an appropriate monitoring network that would work in concert with such a 7 standard to result in the intended degree of protection. As discussed below in section IV.B, the 8 evaluation of monitoring methods focuses on the dev elopment of Federal Reference 9 Methods/Federal Equivalent Methods (FRM/FEM) for NOy and SOx. EPA notes that the 10 monitoring program described here is intended to supplement existing monitoring programs and 11 will require additional resources

. 12 A. Field Pilot Program 13 This section presents the objectives of a field pilot program (section IV.A.1) that would 14 gather relevant field data over a five

- ye ar period in a sample of 3 to 5 sensitive ecoregions across 15 the country . An overview of the scope and structure of the field program, with a focus on 16 measurements of ambient air indicators of oxides of nitrogen and sulfur , is presented in section 17 IV .A.2. Section IV.A.3 explains the role of additional complementary measurements beyond the 18 ambient air indic ators that would be included in the program, and section IV.A.4 discusses a 19 parallel longer

- term research agenda, both of which are guided b y the uncertainties discussed 20 above in section III. Section IV.A.5 identifies implementation challenges presented b y an AAI
21 based standard that could be addressed in parallel with a field pilot program. Section IV.A.6 22 discusses engagement with stakeholder groups as part of the planned pilot program. 23 1. Objectives 24Draft

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193 Consideration of a new multi

- pollutant standard to addr ess deposition

- related effects on 1 sensitive aquatic ecoregions raises unique challenges relative to those typically raised in reviews 2 of existing NAAQS for which an established network of FRM/FEMs, designed to measure the 3 indicator pollutant , is generall y available. The primary goal of this field pilot program, and the 4 related monitoring program discussed in section IV.B, is to enhance our understanding of the 5 degree of protectiveness that would likely be afforded by

a standard based on the AAI, as 6 describ ed above in section III, through the following objectives: 7 (1) evaluate measurement methods for the ambient air indicator s of NOy and SOx and 8 consider designation of such methods as FRMs; 9 (2) examine the variability and improve characterization of concentration and 10 deposition patterns of oxides of nitrogen and sulfur, as well as reduced forms of 11 nitrogen, within and across a number of sensitive ecoregions across the country ; 12 (3) develop updated ecoregion

- specific factors (i.e., F1 through F4) for the AAI 13 equation based in part on new observed air quality data within the sample ecoregions 14 as well as on updated nationwide air quality model results and expanded critical load 15 data bases , and explore alternative approaches for developing such representat ive 16 factors; 17 (4) calculate ecoregion

- specific AAI values using observed NOy and SOx data and 18 updated ecoregion

- specific factors to examine the extent to which the sample 19 ecoregions would meet a set of alternative AAI

- based standards; 20 (5) develop air moni toring network design criteria for an AAI

- based standard; 21 (6) assess the use of total nitrate measurements as a potential alternative indicator for 22 oxides of nitrogen; 23Draft

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194 (7) supoprt related longer

- term research efforts, including enhancements to and 1 evaluation of modeled dry deposition algorithms; and 2 (8) facilitate stakeholder engagement in addressing implementation issues associated 3 with possible future adoption of an AAI

- based st andard. 4 2. Overview of Field Pilot Program 5 The CASTNET program (Figure IV
- 1) affords an available infrastructure relevant to an 6 AAI
- based standard, given the location of sites in some acid

- sensitive ecoregions and various 7 measurements of sulfur and nitrog en species. The EPA plans to use CASTNET sites in selected 8 acid

- sensitive ecoregions to serve as the platform for this pilot program, potentially starting in 9 late 2012 and extending through 2018. CASTNET sites in three to five ecoregions in acid

10 sensitiv e areas would collect NOy and SOx (i.e., SO 2 and p

- SO 4 ) measurements over a five

- ye ar 11 period . The initial step in developing a data base of observed ambient air indicators for oxides of 12 nitrogen and sulfur requires the addition of NOy samplers at the pilo t study sites so that a full 13 complement of indicator measurements are available to calculate AAIvalues . These CASTNET 14 sites would also be used to make supplemental observations useful for evaluation of CMAQ’s 15 characterization of factors F2

–F4 in the AAI equation. 16 The selected ecoregions would account for geographic variabilit y b y including regions 17 from across the U.S., including the east, upper midwest and west. Each selected region would 18 have at least two existing CASTNET sites. Each of the pilot CASTNET sites would be used to 19 verify

the performance of the established methods, data retrieval and reporting procedures, and 20 application of the measurements into the AAI equation. 21 22Draft

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195 1 Figure IV - 1. Location of CASTNET sites in relation to acid sensitive ecoregions. 2
3 Over the course of this five

- year pilot program, the most current national air quality 4 modeling , based on the most current national emissions inventory , would be used to develop an 5 updated set of F 2

– F4 factors. A parallel multi

- agency national critical load data base 6 development effort would be used as the basis for calculating updated F1 factors. As discussed 7 above in section III.B, these factors would be based on av erage parameter values across an 8 ecoregion. Using this new set of F factors, observations of NOy and SOx derived from the pilot 9 program , averaged across each ecoregion, would be used to calculate AAI values in the sample 10 ecoregions

. The data from the pi lot program would also be used to examine alternative 11Draft

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196 approaches to generating representative air quality values, such as examining the appropriateness 1 of spatial averaging in areas of high spatial variability

. 2 3 . Complementary Measurements 3 Complementary measurements may be performed at some sites in the pilot network to 4 reduce uncertainties in the recommended methods and better characterize model performance 5 and application to the AAI. The CASAC Air Monitoring and Methods Subcommittee (AMMS) 6 advised EPA that such supplemental measurements were of critical importance in a field 7 measurement program related to an AAI
- based standard (Russell and Samet, 2011 b ). 8 Candidate complementary measurements to address sulfur, in addition to those provided 9 by the CASTNET filter pack (CFP), include trace gas continuous SO 2 and speciated PM 2.5 10 measurements. The co

- located deploy ment of a continuous SO 2 analy zer with the CFP SO 2 will 11 provide test data for determining suitability of continuous SO 2 measurements as an FE M, as well 12 as producing valuable time series data for model evaluation purposes. The weekl y averaging 13 time provided by

the CFP adequatel y address es the annual

- average basis of an AAI
- based 14 secondary standard, but would not be applicable to short

- term (i.e ., 1
- hour) average associated 15 with the primary

SO 2 standard. Conversel y, because of the low concentrations associated with 16 many

acid

- sensitive ecoregions, existing SO 2 FRMs designated for use in determining 17 compliance with the primary standard would not n ecessarily be appropriate for use in 18 conjunction with an AAI

- based secondary standard

. 19 Co
- locating the PM 2.5 sampler used in the EPA Chemical Speciation Network and the 20 IMPROVE network at pilot network sites would allow for characterizing the relationsh ip 21 between the CFP

- derived p - SO 4 and the speciation samplers used throughout the state and local 22 air quality

networks. Note that CASTNET already has several co

- located IMPROVE chemical 23Draft

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197 speciation samplers. Because the AAI equation is based on concentrati on of p - SO 4 , the original 1 motivation for capturing all particle size fractions is not as important relative to simply capturing 2 the concentration of total p - SO 4 . 3 Candidate measurements to complement oxidized nitrogen measurements, in addition to 4 the CFP , include a mix of continuous and periodic sampling for the dominant NOy species, 5 namely

NO, true NO 2 , PAN, HNO 3 , and p - NO 3 . While there are several approaches to acquiring 6 these measurements, perhaps the most efficient strategy would take advantage of the available 7 CFP total nitrate, and add a three

- channel chemiluminescence instrument that will cycle between 8 NOy , true NO 2 an d NO by adding photol yt ic detection for true NO 2 . Other options for 9 measuring true NO 2 would include adding either a stand

- alone photol yt ic or cavit y ring

- down 10 spectroscop y instrument. Measurements of PAN may be acquired either on a periodic basis 11 through canister sampling and subsequent laboratory anal ys is or through emerging in

- situ 12 sampling and anal ys is methods. Although the CFP yields a reliable measurement of total nitrate, 13 t - NO 3 (i.e., the sum of HNO 3 and p

- NH 4 ) value, strong consideration may be g iven to direct 14 measurement of HNO 3 , which has the highest deposition velocity of all the dominant NOy 15 species. Similar to the use of continuous SO 2 data, these speciated NOy data serve two purposes 16 by evaluating total NOy

instrument behavior and air quali ty model evaluation. The measurement 17 of individual NOy

species can be used to generate site

- specific NOy values for comparison to 18 modeled NOy , and will likely provide insight into and improvement of modeled dry deposition. 19 The CASAC AMMS (Russell and Same t, 2011 b ) recommended that EPA consider the 20 use of t - NO 3 obtained from CASTNET sampling as an indicator for NOy , reasoning that t - NO 3 is 21 ty picall y a significant fraction of deposited oxidized nitrogen in rural environments and 22 CASTNET measurements are wide ly available. Collection of this data would support further 23Draft

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198 consideration of using the CFP for t - NO 3 as the indicator of oxides of nitrogen for use in an AAI

1 based secondary

standard. 2 The CASAC AMMS also recommended that total NHx (NH 3 and p

- NH 4 ) be cons idered 3 as a prox y for reduced nitrogen species, reasoning that the subsequent partitioning to NH 3 and p

4 NH 4 may

be estimated using equilibrium chemistry calculations. Reduced nitrogen 5 measurements are used to evaluate air quality modeling which is used in generating factor F2. 6 Additional studies are needed to determine the applicability of NHx measurements and 7 calculated values of NH 3 and NH 4 to the AAI. 8 The additional supplemental measurements of speciated NOy , continuous SO 2 and NHx 9 will be used in future air quality modeling evaluation efforts. Because there often is significant 10 lag in the availability of contemporary

emissions data to drive air quality modeling, the complete 11 use of these data sets will extend beyond the five

- year collection period of the pilot program. 12 Consequently , the immediate application of those data will address instrument performance 13 comparisons that explore the feasibility of using continuous SO 2 instruments in rural 14 environments, and using the speciated NOy data to assess NOy instrument performance. 15 Although contemporary air quality

modeling will lag behind measurement data availability , the 16 observations can be used in deposition models to compare observed transference ratios with the 17 previously

calculated transference ratio s to test temporal stability of the ratios. 18 An extended water quality sampling effort should parallel the air quality measurement 19 program to address some of the uncertainties related to factor F1 and the representativeness of 20 the nth percentile critical load as discussed in section III. B.5.b.i. The objective of the water 21 quality

sampling would be to develop a larger data base of critical loads in each of the pilot 22 ecoregions such that the nth percentile can adequately be characterize d in terms of representing 23Draft

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199 all water bodies. Opportunities to leverage and perhaps enhance existing ecosystem modeling 1 efforts enabling more advanced critical load modeling and improved methods to estimate base 2 cation production would be pursued. For example, areas with ongoing research studies 3 producing data for d ynamic critical load modeling would be considered when selecting the pilot 4 ecoregions. 5 4. Complementary areas of research 6 The EPA recognizes that a source of uncertaint y in an AAI

- based secondary standard that 7 would not be directl y addressed in the pilot program stems from the uncertainty in the model 8 used to link atmospheric concentrations to dry deposition fluxes. Currently, there are no ongoing 9 direct dry

deposition measurement studies at CA STNET sites that can be used to evaluate 10 modeled results. It was strongl y recommended b y CASAC AMMS that a comprehensive 11 sampling

- intensive study be conducted in at least one, preferabl y two sites in different 12 ecoregions to assess characterization of dry deposition of sulfur and nitrogen. These sites would 13 be the same as those for the complementary measurements described above, but they would 14 afford an opportunity to also complement dry

deposition process research that benefits from the 15 ambient air measur ements collected in the pilot program. The concerns regarding uncertainties 16 underlying an AAI

- based secondary standard suggests that research that includes dry deposition 17 measurements and evaluation of dry deposition models should be a high priority

. 18 Simi lar leveraging should be pursued with respect to ecosystem research activities. For 19 example, studies that capture a suite of soil, vegetation, hydrological, and water qualit y 20 properties that can help evaluate more advanced critical load models would compl ement the 21 atmospheric

- based pilot program. In concept, such studies could provide the infrastructure for 22 true multi

- pollutant , multi

- media “super” sites assuming the planning, coordination , and resource 23Draft

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200 facets can be aligned. While this discussion emphas izes the opportunity of leveraging ongoing 1 research efforts, consideration could be given to explicitly including related research components 2 directly

in the pilot program. 3 5. Implementation challenges 4 The CAA requires that once a NAAQS is established, designation and implementation 5 must move forward. With standard as innovative as the AAI

- based standard considered in this 6 review, the Administrator believes that its success will be greatl y improved if, while additional 7 data are being collected to reduce the uncertainties discussed above, the implementing agencies 8 and other stakeholders have an opportunity to discuss and thoroughl y understand how such a 9 standard would work. And since, as noted above, emissions reductions that are directionally 10 correct t o reduce aquatic acidification will be occurring as a result of other CAA programs, the 11 Administrator believes that this period of further discussion will not delay progress but w i ll 12 ensure that once implementation is triggered, agencies will be prepared t o implement it 13 successfull y. 14 Consideration of an AAI

- based secondary standard for oxides of nitrogen and sulfur 15 would present significant implementation challenges because it involves multiple, regionall y 16 dispersed pollutants and relativel y complex compliance determinations. The anticipated 17 implementation challenges fall into three main categories: monitoring and compliance 18 determinations for area designations, pre

- construction permit application analy ses of individual 19 source impacts, and State Implementation Plan (SIP) development. Several overarching 20 implementation questions that we anticipate will be addressed in parallel with the field pilot 21 program’s five

- ye ar data collection period include: 22 (1) What are the appropriate monitoring network density and siting requirements to 23Draft

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201 support a compliance s ys tem based on ecoregions? 1 (2) Given the unique spatial nature of the secondary standard (e.g., ecoregions), what are 2 the appropriate parameters for est ablishing nonattainment areas? 3 (3) How can new or mod ified major sources of oxides of nitrogen and oxides of sulfur 4 emissions assess their ambient impacts on the standard and demonstrate that they are not 5 causing or contributing to a violation of the NAAQS f or preconstruction permitting? 6 (4) What additiona l tools, information, and planning structures are needed to assist states 7 with SIP development, including the assessment of interstate pollutant transport and 8 deposition? 9 (5) Would transportation conformity appl y in nonattainment and maintenance areas for 10 this secondary standard, and, if it does, would satisfying requirements that apply for 11 related primary standards ( e.g., ozone, PM 2.5 , and NO 2 ) be demonstrated to satisfy 12 requirements for this secondary standard? 13 6. Final monitoring plan development and stakeholder participation 14 The existing CASTNET sampling site infrastructure provides an effective means of 15 quickly

and efficiently

deploying a monitoring program to support potential implementation of 16 an AAI
- based secondary standard, and also provides an a dditional opportunity for federall y 17 managed networks to collaborate and support the states, local agencies and Tribes (SLT) in 18 determining compliance with a secondary standard. A collaborative effort would help to 19 optimize limited federal and SL T monitori ng funds and would be beneficial to all involved. 20 CASTNET is alread y a stakeholder

- based program with over 20 participants and contributors, 21 including federal, state and tribal partners. CASTNET partners are willing to adapt the network 22Draft

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202 to meet the monit oring requirements of an AAI

- based secondary standard, provided the current 1 requirements for program assessment continue to be met. 2 The CASAC AMMS generall y endorsed the technical approaches used in CASTNET, but 3 concerns were raised b y i ndividu al representatives of State agencies concerning the perception of 4 EPA
- controlled management aspects of CASTNET and data ownership. Potential approaches to 5 resolve these issues will be developed and evaluated in existing NACAA/EPA ambient air 6 monitoring wor

kgroups. The EPA Office of Air and Radiation (which includes the Office of Air 7 Quality

Planning Standards, OAQPS; and the Office of Atmospheric Program’s Clean Air 8 Markets Division, OAP

- CAMD), and their partners on the NACAA monitor steering committee 9 wil l develop a prioritized specific plan that identifies the 3

- 5 ecoregions and the instrumentation 10 to be deploy ed. EPA anticipates that a cost estimate of the plan with priorities and options will 11 be developed b y J anuary , 2012. Although this pilot program is focused on data collection, the 12 plan will include details of the data anal ys is approaches as well as a vehicle that incorporates 13 engagement from those within EPA and SL Ts to foster progress on the implementation questions 14 noted above in section IV.A. 5. 15 If an AAI

- based secondary standard were to be set in the future, deploy ment of a full 16 national network would follow the pilot monitoring program. The number of sites deploy ed in 17 the network will lead to increased confidence in capturing spatial patterns of air qualit y. 18 Recognizing that this section presents the general elements of the field pilot programs, EPA 19 intends to develop a more detailed field pilot program plan through a process that will engage the 20 air quality

management and research (atmospheri c and ecos ys tem) communities , as well as other 21 federal agencies, state and local agencies , and non - government based centers of expertise. The 22 EPA is seeking comment and input on all aspects of this field pilot program. 23Draft

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203 . B. Evalu ation of Monitoring Methods 1 The EPA generally relies on monitoring methods that have been designated as FRMs or 2 FEMs for the purpose of determining the attainment status of areas with regard to existing 3 NAAQS. Such FRMs or FEMs are generall y required to measure the air quality indi cators that 4 are compared to the level of a standard to assess compliance with a NAAQS. Prior to their 5 designation b y EPA as FRM/FEMs through a rulemaking process, these methods must be 6 determined to be applicable for routine field use and need to have bee n experimentally validated 7 by meeting or exceeding specific accuracy , reproducibility , and reliabilit y criteria established b y 8 EPA for this purpose. As discussed above in section III .A, the ambient air indicators being 9 considered for use in an AAI

- based s tandard include SO 2 , particulate sulfate (p

- SO 4 ), and total 10 reactive oxides of nitrogen (NOy). 11 The CASTNET provides a well established infrastructure that would meet the basic 12 location and measurement requirements of an AAI

- based secondary standard given the rural 13 placement of sites in acid sensitive areas. In addition, CFPs currentl y provide very economical 14 weekl y, integrated average concentration measurements of SO 2 , p - SO 4 , ammonium ion (NH 4 ) 15 and t - NO 3 , the sum of HNO 3 and p - NO 3 . 16 While routinely operated instruments that measure SO 2 , p - SO 4 , NOy and/or t - NO 3 exist, 17 instruments that measure p - SO 4 , NOy , t - NO 3 , or the CFP SO 2 have not been designated b y EPA 18 as FRMs or FEMs The EPA’s ORD has initiated work that will support future FRM 19 designations b y EPA for SO 2 and p

- SO 4 measurements based on the CFP . Such a designation by 20 EPA could be done for the purpose of facilitating consistent research related to an AAI

- based 21 standard and/or in conjunction with setting and supporting an AAI
- based secondary standard. 22Draft

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204 Based on extensive review of literature and available data, the EPA has identified 1 potential methods that appear suitable for measuring each of the three components of the 2 indicators. These three methods are being considered as new FRMs to be use d for measuring the 3 ambient concentrations of the three components that would be needed to determine compliance 4 with an AAI

- based secondary standard. 5 For the SO 2 and p

- SO 4 measurements, EPA is considering the CFP method, which 6 provides weekl y average conce ntration measurements for SO 2 and p

- SO 4 . This method has been 7 used in the EPA’s CASTNET monitoring network for 15 y ears, and strongly indicates that it will 8 meet the requirements for use as an FRM for the SO 2 and p

- SO 4 concentrations for an AAI

- based 9
seco ndary standard. 10 Although the CFP method would provide measurements of both the SO 2 and p

- SO 4 11 components in a unified sampling and anal ys is procedure, individual FRMs will be considered 12 for each. The EPA recognizes that an existing FRM to measure SO 2 conc entrations using ultra

13 violet fluorescence (UVF) exists (40 CFR Part 50, Appendix A

- 1) for the purpose of monitoring 14 compliance for the primary SO 2 NAAQS. However, several factors suggest that the CFP method 15 would be superior to that UVF FRM for monitorin g compliance with an AAI

- based secondary 16 standard and will be discussed in more detail below. 17 For monitoring the NOy component, a continuous anal yz er for measuring NOy is 18 commerciall y available and is considered to be suitable for use as an FRM. This method is 19 similar in design to the existing NO 2 FRM (described in 40 CFR Part 50, Appendix F), which is 20 based on the ozone chemiluminescence measurement technique. The method is adapted to and 21 further optimized to measure all oxides of nitrogen (NOy). Ho wever, this NOy method requires 22 further evaluation before it can be full y confirmed as a suitable FRM. The EPA is currentl y 23Draft

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205 completing a full scientific assessment of the NOy method to determine whether it would be 1 appropriate to consider for designation by EPA as an FRM. Specific details on these three 2 methods are given below. 3 On February 16, 2011, EPA presented this set of potential FRMs to the CASAC AMMS 4 for their consideration and comment. In response, the CASAC AMMS stated that, overall, it 5 believes that EPA’s planned evaluation of methods for measuring NOy , SO 2 and p

- SO 4 as 6 ambient air indicators are suitable approaches in concept. On supporting the CFP method as a 7 potential FRM for SO 2 , CASAC stated that they felt that the CFP is adequate for measuring 8 long

- term average SO 2 gas concentrations in rural areas with low levels (less than 5 ppbv) and is 9 therefore suitable for consi deration as an FRM. For p

- SO 4 , CASAC generally supports the use 10 of the CFP as a potential FRM for measuring p

- SO 4 for an AAI

- based secondary standard. The 11 method has been relatively well

- characterized and evaluated, and it has a documented, long

- term 12 tra ck record of successful use in a field network designed to assess spatial patterns and long

- term 13 trends. 14 On supporting the photometric NOy method as a potential FRM, CASAC concluded that 15 the existing NOy

method is generall y an appropriate approach for the indicator. However, 16 CASAC agrees that additional characterization and research is needed to fully understand the 17 method in order to designate it as a FRM. The EPA is now soliciting public comment on these 18 methods as to their adequacy , suitability , and re lative merits as FRMs for purposes of monitoring 19 to determine compliance with an AAI

- based secondary standard. 20 1. Potential FRMs for SO2 and p

- SO 4 21 The CFP is a combined, integrated sampling and analy sis method based on the well

22 established measurement tec hnology that has been used extensively in EPA’s CASTNET 23Draft

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206 monitoring network (see http://www.epa.gov/castnet ). This method is in current use at over 80 1 monitoring sites and has been in use at not less than 40 sites f or over 15 ye ars. This method 2 employ s a relativel y simple and inexpensive sampler and uses four 47

- mm filters placed in an 3 open - faced filter pack to simultaneously collect integrated filter samples for the SO 2 and p

- SO 4 4 components. In addition, the CFP i s also capable of the collection of t

- NO 3 , the sum of HNO 3 5 and p - NO 3 . 6 The first stage of the filter pack assembly contains a Teflon® filter that collects p - SO 4 2
- 7 and p - NO 3 , the second stage contains a n yl on filter that collects SO 2 (as SO 4 2
- ) and HNO 3 , and 8 the third stage contains two cellulose fiber filters impregnated with potassium carbonate (K 2 CO 3 ) 9 that collect an y remaining SO 2 (as SO 4 2
- ). The sampler collects one

- week integrated samples at a 10 very

low, controlled flow rate (1.5 or 3 L/min) in an attempt simulate actual deposition. Weekly 11 averaged SO 2 and p

- SO 4 concentrations could then be averaged over a 1

- year period to calculate 12 annual average values. 13 Upon sample completion, the species

- specific filters are extracted, with subsequent 14 anal ys is b y the well

- established and documented ion chromatographic (IC) anal yt ical technique. 15 During the IC anal ys is, an aliquot of a filter extract is injected into a stream of eluent (ion 16 chromatograph y mobile phase, generall y a millimolar

- strength solution of carbonate

17 bicarbonate) and passed through a series of ion exchangers. The anions of interest are separated 18 on the basis of their relative affinities for a low capacity and the strongl y basic anion exchanger 19 (guard and separator column). The separated anion s are directed onto a cation exchanger 20 (suppressor column) where they are converted to their highl y conductive acid form, and the 21 eluent is converted to a weakl y conductive form. The now

- separated anions, each in their acid 22 form, are measured b y conductiv ity . They are identified on the basis of retention time compared 23Draft

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207 to that of standards and quantified by measurement of peak area compared to the peak areas of 1 calibration standards. 2 Calibration and quality assurance for the method are applied to the s ample filters, the 3 anal yt ical processes, and the flow rate measurement and control aspects of the sampler. Overall 4 method performance is ty pically assessed with collocated samplers. These quality assurance 5 techniques are routinel y used and have proved ad equate for other t ypes of FRMs and equivalent 6 methods in air monitoring network service. 7 The measurement and analy tical procedures and past performance data associated with 8 the C FP method are well documented and available through Quality Assurance Performance 9 Plans (QAPPs), Standard Operating Procedures ( SOPs ) and annual reports (US EPA, 2010a and 10 2010b). The accumulated database on the CFP method is substantial and indicates that the 11 method is sound, stable and has good reliability in routine, field operation. Data qualit y 12 assessment results show the method to have good reproducibility , with collocated and anal yt ical 13 precision values in the range of 2%

– 10% (excluding very low concentration measurements near 14 the method detection limits; US EPA 2010b). 15 Data qualit y objectives (DQOs) for a new FRM would be based upon current DQOs 16 being used for this method by EPA’s OAP/CAMD and the National Park Service (NPS), the 17 federal managers of CASTNET (US EPA, 2010a). In its current state, the CFP method is 18 expected to meet or exceed (as past CASTNET data have indicated; US EPA, 2010b) the 19 expected FRM DQOs, even when deplo ye d in new monitoring networks outside of CASTNET. 20 In addition, CASTNET samples have agreed favorably with other measures of SO 2 and p

- SO 4 in 21
comparison studies. For example, in direct comparison with an annular denuder sampler (ADS) 22 method, CASTNET/ADS ratios for SO 2 and p

- SO 4 were generall y on the order of 0.9

- 1.1 23Draft

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208 (Lavery et al, 2009; Sickles et al, 1999; Sickles et al, 2008) , thus illustrating the accuracy of the 1 CFP method in the determination of long

- term average SO 2 and p

- SO 4 concentrations. The EPA 2 believes that the CFP method would be fully adequate as an FRM in determining y earl y average 3 SO 2 and p

- SO 4 concentrations for compliance determination purposes. 4 The EPA recognizes that an existing FRM for SO 2 has proven adequate for the purposes 5 of monitoring compliance for the primary SO 2 NAAQS, specifically the newly

- promulgated 1
- 6 hour standard. However, th is FRM is better suited to the shorter

- term, higher concentration 7 primary

and secondary

SO 2 NAAQS, and there is substantial uncertaint y as to the adequacy of 8 this SO 2 FRM for monitoring the lower concentrations relevant to determining compliance with 9 an AA I
- based secondary standard. The performance specifications for SO 2 FRM anal yz ers (40 10 CFR Part 53, Table B

- 1) require a lower detectable limit (LDL) of 0.002 ppm for the standard 11 measurement range and 0.001 ppm for the lower measurement range. These requi rements 12 correspond to mass per unit volume concentrations of 5.24 and 2.62 μg/m 3 , respectivel y. 13 Analy sis of 2009 CASTNET data shows that of the 84 CASTNET sampling sites, 63 measured 14 annual average SO 2 concentrations below even the lower of these LDL requirements of 2.62 15 μg/m 3 for the lower range SO 2 FRM (US EPA, 2010a). In addition, 11 of the 84 sites measured 16 annual (2009) average SO 2 concentrations very near or below the manufacturers’ reported 17 detection limits for trace level UVF SO 2 monitors. Furth er, it is likely that the number of sites 18 with annual average SO 2 concentration below both the SO 2 FRM LDL and the manufacturers 19 reported detection limits will increase due to expected declines in mean SO 2 concentrations (US 20 EPA, 2010b). For these reasons, EPA is considering the CFP method for use as the FRM for 21 monitoring the SO 2 component of an ambient air indicator for oxides of sulfur, with a 22Draft

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209 recommendation for additional study and data collection to further evaluate the possible 1 applicability

of the co ntinuous UVF SO 2 FRM for this purpose. 2 2. Potential FRM for NOy 3 Atmospheric concentrations of NOy are measured continuously by an analyzer that 4 photometrically

measures the light intensity , at wavelengths greater than 600 nanometers (nm), 5 resulting from the chemiluminescent reaction of ozone (O 3 ) with NO in sampled ai r. This 6 method is very

similar to the chemiluminescence NO/NO 2 anal yz ers widely used to collect NO 2 7 monitoring data for determining compliance with the NO 2 NAAQS. The various oxides of 8 nitrogen species, excluding NO, are first quantitatively reduced to N O b y means of a catal yt ic 9 converter. These species include NO 2 , HNO 2 , PANs, HNO 3 and p - NO 3 . The NO, which 10 commonly

exists in ambient air, passes through the converter unchanged, and, when combined 11 with the NO resulting from the catal yt ic conversion of th e other oxides of nitrogen, a 12 measurement of the total NOy concentration results. To maximize the conversion of the more 13 chemicall y active oxides of nitrogen species, the converter is located externally , at or near the air 14 sample inlet probe. This locati on minimizes losses of these active species that could otherwise 15 occur from chemical reactions and wall losses in the sample inlet line. 16 The NOy

anal yz er is a suitable, commerciall y produced continuous chemiluminescence 17 anal yz er that includes an ozone gene rator, a reaction cell, a photometric detector, wavelength 18 filters as necessary

to reduce sensitivity

to wavelengths below 600 nm, a pump and flow control 19 sy stem to draw atmospheric air through the converter and into the reaction cell, a suitable 20 converter , a s ys tem to control the operation of the anal yz er, and appropriate electronics to 21 process and quantitatively scale the photometric signals. The converter contains a catal ys t such 22 as moly

bdenum and is heated to an optimum temperature designed to optimize the conversion of 23Draft

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210 the various oxides of nitrogen to NO. It is connected to the anal yz er via suitable lengths of 1 Teflon® tubing. Hourly NOy measurements obtained by the anal yz er would be averaged over 2 the same 7

- day

period used by

the CFP method to measure the SO 2 and p

- SO 4 components, with 3 further averaging over a 1

- ye ar period

. 4 Commercial NOy analyzers are currentl y available, and the anal yz ers have been used for 5 a variet y of monitoring applications. During the 2006 TexAQS Radical and Aerosol 6 Measurement Project (TRAMP), Luke et. al. , 2010, compared measured NOy concentrations 7 obtained with an NOy instrument based upon the above mentioned methodology with the sum of 8 measured individual NOy species (i.e., NOy i = NO+NO 2 +HNO 3 +PANs+HNO 2 +p - NO 3 ). Th is 9 comparison yi elded excellent overall agreement during both day ([NOy ](ppb) = [NOy i ](ppb) × 10 1.03

– 0.42; r2 = 0.9933) and night time ([NOy](ppb) = [NOy i ](ppb) × 1.01

– 0.18; r2 = 0.9975) 11 periods ( Luke et al , 2010 ). The results of this study show that this NOy method is capable of the 12 accurate determination of all the atmosphericall y relevant NOy components, resulting in an 13 accurate determination of total NOy concentrations. NOy

instruments have been routinel y 14 op erated in networks such as SEARCH, dating back several years. In addition, state monitoring 15 agencies across the U.S. have begun, starting in 2009, the routine operation of commercially 16 available NOy

instrumentation in anticipation of EPA’s NCore network transitioning to full 17 operation in 2011. 18 These initial assessments described above are promising and indicate that the photometric 19 NOy

method appears to be accurate, reliable, and capable of routine network operation. As a 20 result, the method is most like ly capable for use as an FRM for determining atmospheric NOy 21 concentrations as a component in determining compliance with an AAI

- based secondary 22 standard. However, as described below, this continuous method for NOy requires additional 23Draft

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211 time for further eva luation before it can be full y confirmed for adoption as a FRM. The EPA has 1 identified measurement uncertainties and some remaining science questions associated with this 2 method. Among these are (a) the ability of the method to capture all components of NOy 3 relevant to nitrogen deposition, (b) the efficiency of the moly

bdenum converter in converting all 4 oxides of nitrogen to NO for detection (excluding NO 2 , as this conversion is alread y well 5 documented), (c) appropriate inlet height specifications to mini mize any bias associated with 6 vertical concentration gradients of key NOy components, (d) identification and quantification of 7 potential measurement interferences in the NOy determination, and (e) development and 8 demonstration of effective calibration/chal lenge procedures to best represent the various 9 mixtures of NOy

components that are expected to be present in the different air sheds across the 10 U.S. 11 To address these NOy method uncertainties and to fully assess this method for use as the 12 NOy

FRM, EPA has d eveloped a detailed research plan ( Russell and Samet, 2011b) which was 13 presented to the CASAC AMMS on February 16, 2011. In response, CASAC recognized the 14 need for, and supported the general outline of EPA’s research plan to evaluate the NOy method 15 for po tential designation as an FRM (US EPA, 2011). In addition, the CASAC AMMS 16 suggested additional areas of research associated with the photometric NOy method that warrant 17 further assessment prior to final designation of the method as the NOy FRM. These inc lude 18 operation of the method during extremel y low temperature conditions to investigate possible 19 condensation in sample lines, method detection limits relative to low levels expected in remote 20 areas, and ambient

- based method evaluations in various air shed s across the U.S. In response to 21 these CASAC AMMS suggestions, EPA is carrying out studies, in addition to the tasks outlined 22 in the research plan, for the NOy method. The results of these studies will likely take a y ear or 23Draft

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212 more to become available. As noted previousl y, EPA anticipates that these results will be 1 favorable and will confirm the adequacy of the NOy

method as a suitable FRM for determining 2 compliance with an AAI

- based secondary standard. 3 V. Statutory and Executive Order Reviews 4 A. Executive Order 12866: Regulatory Planning and Review and Review and Executive 5 Order 13563: Improving Regulation and Regulatory Review 6 Under Executive Order 12866 (58 FR 51735, October 4, 1993), this action is a 7

“ significant regulatory action” because it was deemed to “raise novel legal or policy issues.” 8 Accordingl y, EPA submitted this action to the Office of Management and Budget (OMB) for 9 review under Executive Orders 12866 and 13563 (76 FR 3821, January 21, 2011), and an y 10 changes made in response to OMB r ecommendations have been documented in the docket for 11 this action. 12 B. Paperwork Reduction Act 13 This action does not impose an information collection burden under the provisions of the 14 Paperwork Reduction Act, 44 U.S.C. 3501 et seq. Burden is defined at 5 CFR 1320.3(b). There 15 are no information collection requirements directly associated with the establishment of a 16 NAAQS under section 109 of the CAA. 17 C. Regulatory Flexibility Act 18 For purposes of assessing the impacts of today ’s rule on small entities, smal l entity is 19 defined as: (1) a small business that is a small industrial entity as defined by

the Small Business 20 Administration’s (SBA) regulations at 13 CFR 121.201; (2) a small governmental jurisdiction 21 that is a government of a city , count y, town, scho ol district or special district with a population of 22Draft

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213 less than 50,000; and (3) a small organization that is any not - for

- profit enterprise which is 1 independentl y owned and operated and is not dominant in its field. 2 After considering the economic impacts of today 's proposed rule on small entities, I 3 certify

that this action will not have a significant economic impact on a substantial number of 4 small entities. This proposed rule will not impose any requirements on small entities. Rather, 5 this rule establish es national standards for allowable concentrations of oxides of nitrogen and 6 sulfur in ambient air as required b y section 109 of the CAA. See also American Trucking 7 Associations v. EPA. 175 F. 3d at 1044

- 45 (NAAQS do not have significant impacts upon smal l 8 entities because NAAQS themselves impose no regulations upon small entities). We continue to 9 be interested in the potential impacts of the proposed rule on small entities and welcome 10 comments on issues related to such impacts 11 D. Unfunded Mandates Reform Act 12 Title II of the Unfunded Mandates Reform Act of 1995 (UMRA), Public Law 104

- 4, 13 establishes requirements for Federal agencies to assess the effects of their regulatory actions on 14 State, local, and Tribal governments and the private sector. Under sect ion 202 of the UMRA, 15 EPA generall y must prepare a written statement, including a cost

- benefit analy sis, for proposed 16 and final rules with “Federal mandates” that may result in expenditures to State, local, and Tribal 17 governments, in the aggregate, or to th e private sector, of $100 million or more in any 1 y ear. 18 Before promulgating an EPA rule for which a written statement is needed, section 205 of the 19 UMRA generally

requires EPA to identify and consider a reasonable number of regulatory 20 alternatives and to adopt the least costly , most cost - effective or least burdensome alternative that 21 achieves the objectives of the rule. The provisions of section 205 do not apply when they

are 22 inconsistent with applicable law. Moreover, section 205 allows EPA to adopt an alternative 23Draft

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214 other than the least costly, most cost

- effective or least burdensome alternative if the 1 Administrator publishes with the final rule an explanation why that alternative was not adopted. 2 Before EPA establishes any regulatory

requirements that may significantl y or uniquely affect 3 small governments, including Tribal governments, it must have developed under section 203 of 4 the UMRA a small government agency plan. The plan must provide for notifying potentiall y 5 affected small governments, enabling off icials of affected small governments to have meaningful 6 and timely

input in the development of EPA regulatory proposals with significant Federal 7 intergovernmental mandates, and informing, educating, and advising small governments on 8 compliance with the regulatory requirements. 9 This action contains no Federal mandates under the provisions of Title II of the 10 Unfundated Mandates Reform Act of 1995 (UMRA), 2 U.S.C. 1531

- 1538 for State, local, or 11 tribal governments or the private sector. Therefore, this actio n is not subject to the requirements 12 of sections 202 or 205. Furthermore, as indicated previously , in setting a NAAQS EPA cannot 13 consider the economic or technological feasibility of attaining ambient air quality standards; 14 although such factors may be co nsidered to a degree in the development of State plans to 15 implement the standards. See also American Trucking Associations v. EPA, 175 F. 3d at 1043 16 (noting that because EPA is precluded from considering costs of implementation in establishing 17 NAAQS, prep aration of a Regulatory Impact Analy sis pursuant to the Unfunded Mandates 18 Reform Act would not furnish any information which the court could consider in reviewing the 19 NAAQS). Accordingl y, EPA has determined that the provisions of sections 202, 203, and 20 5 of 20 the UMRA do not apply to this proposed decision. The EPA acknowledges, however, that any 21 corresponding revisions to associated SIP requirements and air qualit y surveillance requirements, 22 40 CFR part 51 and 40 CFR part 58, respectivel y, might result i n such effects. Accordingl y, 23Draft

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215 EPA will address, as appropriate, unfunded mandates if and when it proposes an y revisions to 40 1 CFR parts 51 or 58. 2 E. Executive Order 13132: Federalism 3 This proposed rule does not have federalism implications. It will not h ave substantial 4 direct effects on the States, on the relationship between the national government and the States, 5 or on the distribution of power and responsibilities among the various levels of government, as 6 specified in Executive Order 13132 because it does not contain legally binding requirements. 7 Thus, the requirements of Executive Order 13132 do not apply to this rule. 8 EPA believes, however, that this proposed rule may be of significant interest to State 9 governments. However, as also noted in sectio n E (above) on UMRA, EPA recognizes that 10 States will have a substantial interest in this rule and any corresponding revisions to associated 11 SIP requirements and air quality surveillance requirements, 40 CFR part 51 and 40 CFR part 58, 12 respectivel y. Theref ore, in the spirit of Executive Order 13132 and consistent with EPA policy 13 to promote communications between EPA and State and local governments, EPA specificall y 14 solicits comment on this proposed rule from State and local officials. 15 F. Executive Order 131 75: Consultation and Coordination with Indian Tribal Governments 16 Executive Order 13175, entitled “Consultation and Coordination with Indian Tribal 17 Governments” (65 FR 67249, November 9, 2000), requires EPA to develop an accountable 18 process to ensure “mean ingful and timely input by tribal officials in the development of 19 regulatory

policies that have tribal implications.” This rule concerns the establishment of 20 national standards to address the public welfare effects of o xides of n itrogen and s ulfur

. 21 This action does not have Tribal implications, as specified in Executive Order 13175 (65 22 FR 67249, November 9, 2000). It does not have a substantial direct effect on one or more Indian 23Draft

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216 tribes, since tribes are not obligated to adopt or implement any NAAQS. Thus, Executive Order 1 13175 does not apply to this rule. 2 G. Executive Order 13045: Protection of Children from Environmental Health &Safety 3 Risks 4 The EPA interprets EO 13045 (62 FR 19885, April 23, 1997) as appl yi ng only to those 5 regulatory

actions that concern health or safet y risks, such that the anal ys is required under 6 section 5

- 501 of the EO has the potential to influence the regulation. This action is not subject to 7 EO 13045 because it addresses the effects on public welfare of oxides of nitrogen an d sulfur and 8 does not establish an environmental standard intended to mitigate health or safet y risks

. 9 H. Executive Order 13211: Actions that Significantly Affect Energy Supply, Distribution or 10 Use 11 This action is not a “significant energy action” as defin ed in Executive Order 13211 (66 12 FR 28355, May

22, 2001), because it is not likely to have a significant adverse effect on the 13 supply , distribution, or use of energy

. This action concerns the establishment of national 14 standards to address the public welfar e effects of oxides of nitrogen and sulfur. This action does 15 not prescribe specific pollution control strategies by which these ambient standards will be met. 16 Such strategies will be developed b y States on a case
- by

- case basis, and EPA cannot predict 17 wh ether the control options selected b y States will include regulations on energy suppliers, 18 distributors, or users. 19 I. National Technology Transfer and Advancement Act 20 Section 12(d) of the National Technology Transfer and Advancement Act of 1995 21 (NTTAA), P ublic Law No. 104

- 113, 12(d) (15 U.S.C. 272 note) directs EPA to use voluntary 22 consensus standards in its regulatory activities unless to do so would be inconsistent with 23Draft

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217 applicable law or otherwise impractical. Voluntary consensus standards are technical standards 1 (e.g., materials specifications, test methods, sampling procedures, and business practices) that are 2 developed or adopted b y voluntary consensus standards bodies. The NTTAA directs EPA to 3 provide Congress, through OMB, explanations when the Age ncy decides not to use available and 4 applicable voluntary consensus standards. 5 This proposed rulemaking does not involve technical standards. Therefore, EPA is not 6 considering the use of any voluntary

consensus standards. 7 J. Executive Order 12898: Federal Actions to Address Environmental Justice in Minority 8 Populations and Low

-Income Populations 9 Executive Order 12898 (59 FR 7629 (Feb. 16, 1994)) establishes federal executive policy 10 on environmental justice. Its main provision directs federal agencies, to the greatest extent 11 practicable and permitted by law, to make environmental justice part of their mission by 12 identifying and addressing, as appropriate, disproportionately high and adverse human health or 13 environmental effects of their programs, policies, and activities on minority populations and low

14 income populations in the United States. 15 EPA has determined that this proposed rule will not have disproportionatel y high and 16 adverse human health or environmental effects on minority or low

- income populat ions because it 17 retain s the level of environmental protection for all affected populations without having any 18 disproportionately

high and adverse human health or environmental effects on any population, 19 including an y minorit y or low

- income population. 20 21 Ref

erences 22 Baker JP; Bernard DP; Christensen SW; Sale MJ. 1990. Biological effects of changes in surface w ater acid

- base 23 chemistry. (State of science / technology report #13).Washington DC: National Acid Precipitation 24 Assessment Program (NAPAP). 25Draft

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218 Bulger AJ; Cosby BJ; Webb JR. 2000. Current, reconstructed past, and projected future status of brook trout 1 ( Salvelinus fontinalis ) streams in Virginia. Can J Fish Aquat Sci, 57, 1515

- 1523. 2 Banzhaf, S., D. Burtraw , D. Evans, and A. Krupnick. 2006. “Valuation of Natur al Resource Improvements in the 3 Adirondacks.” Land Economics 82:445
- 464. 4 Lavery, T.F. C.M. Rogers, R. Baumgardner, and K.P. Mishoe. 2009. Intercomparison of Clean Air Status and 5 Trends Network Nitrate and Nitric Acid Measurements with Data from Other Monit oring Programs . J . Air 6 & Waste Manage. Assoc. , 59, 214
- 226. 7 Lien L; Raddum GG; Fjellheim A. 1992. Critical loads for surface waters: invertebrates and fish. (Acid rain research 8 report no 21). Oslo, Norw ay : Norw egian Institute for Water Research. 9 Luke, W.T. , P.K. Barry, L. Lefer, J. Flynn, B. Rappenglück, M. Leuchner, J.E. Dibb, L.D. Ziemba, C.H. Anderson, 10 and M. Buhr. 2010. Measurements of primary trace gases and NO Y com position in Houston, Texas. 11 Atmospheric Environment , 44, 4068
- 4080. 12 MacAvoy SW; Bulger AJ. 1995. Survival of brook trout (Salvelinus fontinalis) embryos and fry in streams of 13 different acid sensitivity in Shenandoah National Park, USA. Water Air Soil Pollut, 85, 445

- 450. 14 McNulty SG; Cohen EC; Myers JAM; Sullivan TJ; Li H. 2007. Estimates of critical acid loads and exceedances for 15 forest soils across the conterminous United States. Environ Pollut, 149, 281

- 292. 16 NAP AP. 1990. Acid Deposition: State of Science and Technology. National Acid Precipitation Assessment 17 Program. Office of the Director, Washington, DC. 18 NAP AP. (2005). National acid precipitation assessment program report to Congress: An integrated assessment. 19 http://www.esrl.noaa.gov/csd/aqrs/reports/napapreport05.pdf. Silver Spring, MD: National Acid 20 Precipitation Assessment Program (NAPAP); Committee on Environment and Natural Resources (CENR) 21 of the National Science and Technology Council (NSTC). 22 NRC (National Research Council). 2004 . Air quality management in the United States. Washington, DC: National 23 Research Council (NRC); The Nati onal Academies Press. 24 Russell, A and J. M. Samet, 2010a. Review of the Policy Assessment for the Review of the Secondary National 25 Ambient Air Quality Standard for NOx and SOx: First Draft. EPA

- CASAC
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- CASAC
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- 003. 28 Russell, A and J. M. Samet, 2011. Review of the Policy Assessment for the Revie w of the Secondary National 29 Ambient Air Quality Standard for NOx and SOx: FINAL. EPA

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- 005. 30 Russell and Samet, 2011b Review of EPA Draft Documents on Monitoring and Methods for Oxides of Nitrogen 31 (NOx) and Sulfur (SOx) http://yosemite.epa.gov/sab/sabpeople.nsf/WebCommittees/CASAC 32 Sickles II, J.E. , L. L. Hodson, and L. M. Vorburger. 1999
. Evaluation of the filter pack for long

- duration sampling 33 of ambient air

. Atmospheric Environment , 33, 2187
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. Atmospheric Environment , 42, 2062
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219 Stoddard J; Kahl JS; Deviney FA; DeWalle DR; Driscoll CT; Herlihy AT; Kellogg JH; Murdoch PS; Webb JR; 1 Webster KE. (2003). Response of surface water chemistry to the Clean Air Act Amendments of 1990 (No. 2 EPA 620/R

- 03/001). Research Triangle Park, NC; National Health and Environmental Effects Research 3 Laboratory; Office of Research and Development; U.S. Environmental Protection Agency. 4 Sullivan TJ; Driscoll CT; Cosby BJ ; Fernandez IJ; Herlihy AT; Zhai J; Stemberger R; Snyder KU; Sutherland JW; 5 Nierzwicki

- Bauer SA; Boylen CW; McDonnell TC; Nowicki NA. 2006. Assessment of the extent to which 6 intensively studied lakes are representative of the Adirondack Mountain region. (F inal Report no 06

- 7 17).Corvallis, OR; prepared by Environmental Chemistry, Inc. for: Albany, NY; Environmental Monitoring 8 Evaluation and Protection Program of the New York State Energy Research and Development Authority 9 (NYSERDA). 10 US EPA, 1973. “Effects of Sulfur Oxide in the Atmosphere on Vegetation”. Revised Chapter 5 of Air Quality 11 Criteria For Sulfur Oxides. U.S. Environmental Protection Agency. Research Triangle Park, N.C. EPA

- R3
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- 007. U.S. Environmental Protection 15 Agency, Office of Air Quality Planning and Standards, Research Triangle Park, NC. 16 US EPA, 1984a. The Acidic Depositio n Phenomenon and Its Effects: Critical Assessment Review Papers. Volume I 17 Atmospheric Sciences. EPA

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- 016AF. Office of Research and Development, Washington, DC 18 US EPA, 1984b. The Acidic Deposition Phenomenon and Its Effects: Critical Assessment Revi ew Papers. Volume 19 II Effects Sciences. EPA

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- 016BF. Office of Research and Development, Washington, DC 20 US EPA, 1985. The Acidic Deposition Phenomenon and Its Effects: Critical Assessment Document. EPA

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- 21 85/001. Office of Research and Development , Washington, DC 22 US EPA. 1995a. Review of the National Ambient Air Quality Standards for Nitrogen Dioxide: Assessment of 23 Scientific and Technical Information . OAQPS Staff Paper. EPA

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- 005. U.S. Environmental 24 Protection Agency, Office of Air Quality Planning and Standards, Research Triangle Park, NC. September. 25 US EPA. 1995b. Acid Deposition Standard Feasibility Study Report to Congress. U.S. Environmental Protection 26 Agency, Washington, DC. EPA

- 430/R

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- 001a. 27 US EPA 2007. Integrated Review Plan for th e Secondary National Ambient Air Quality Standards for Nitrogen 28 Dioxide and Sulfur Dioxide . U.S. Environmental Protection Agency, Research Triangle Park, NC, EPA

29 452/R

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- 006. 30 US EPA 2008. Integrated Science Assessment (ISA) for Oxides of Nitrogen and Sulf ur Ecological Criteria (Final 31 Report). U.S. Environmental Protection Agency, Washington, D.C., EPA/600/R

- 08/082F, 2008. 32 US EPA 2009. Risk and Exposure Assessment for Revie w of the Secondary National Ambient Air Quality 33 Standards for Oxides of Nitrogen and Oxides of Sulfur

- Main Content

- Final Report. U.S. Environmental 34 Protection Agency, Washington, D.C., EPA

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- 008a 35 US EPA, 2010a. CASTNET Quality Assurance Project Plan, Revision 7.0, October 2010, 36 http://jav a.epa.gov/castnet/ 37 US EPA, 2010b. CASTNET Annual Reports, 2004

- 2009, http://java.epa.gov/castnet/ 38 US EPA 2011. Policy Assessment for the Review of the Secondary National Ambient Air Quality Standards for 39 Oxides of Nitrogen and Oxides of Sulfur. U.S. Environmental Protection Agency, Washington, D.C., EPA

40 452/R

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- 005a. 41Draft

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220 US EPA, 2011b. Federal Reference Methods for NO y and p

- SO 4 for the New Combined NO x and SO x Secondary 1 NAAQS Research Plan, EPA/600/1

- 11/002 January 20, 2011. 2 Wolff, G. T. 1993. CASAC closure letter for the 1993 Criteria Document for Oxides of Nitrogen addressed to U.S. 3 EPA Administrator Carol M. Browner dated September 30, 1993. 4 Wolff, G. T. 1995. CASAC closure letter for the 1995 OAQPS Staff Paper addressed to U.S. EPA Administrator 5 Carol M. Browner dated August 22, 1995. 6