46084 Federal Register / Vol. 76, No. 147 / Monday, August 1, 2011 / Proposed Rules 

ENVIRONMENTAL PROTECTION 

AGENCY 

40 CFR Part 50 

[EPA–HQ–OAR–2007–1145; FRL–9441–2] 

RIN 2060–AO72 

Secondary National Ambient Air 

Quality Standards for Oxides of 

Nitrogen and Sulfur 

AGENCY: Environmental Protection 

Agency (EPA). 

ACTION: Proposed rule. 

SUMMARY: This proposed rule is being 

issued as required by a consent decree 

governing the schedule for completion 

of this review of the air quality criteria 

and the secondary national ambient air 

quality standards (NAAQS) for oxides of 

nitrogen and oxides of sulfur. Based on 

its review, EPA proposes to retain the 

current nitrogen dioxide (NO2) and 

sulfur dioxide (SO2) secondary 

standards to provide requisite 

protection for the direct effects on 

vegetation resulting from exposure to 

gaseous oxides of nitrogen and sulfur in 

the ambient air. Additionally, with 

regard to protection from the deposition 

of oxides of nitrogen and sulfur to 

sensitive aquatic and terrestrial 

ecosystems, including acidification and 

nutrient enrichment effects, EPA is 

proposing to add secondary standards 

identical to the NO2 and SO2 primary 1- 

hour standards and not set a new multipollutant secondary standard in this 

review. The proposed 1-hour secondary 

NO2 standard would be set at a level of 

100 ppb and the proposed 1-hour 

secondary SO2 standard would be set at 

75 ppb. In addition, EPA has decided to 

undertake a field pilot program to gather 

and analyze additional relevant data so 

as to enhance the Agency’s 

understanding of the degree of 

protectiveness that a new multipollutant approach, defined in terms of 

an aquatic acidification index (AAI), 

would afford and to support 

development of an appropriate 

monitoring network for such a standard. 

The EPA solicits comment on the 

framework of such a standard and on 

the design of the field pilot program. 

The EPA will sign a notice of final 

rulemaking for this review no later than 

March 20, 2012. 

DATES: Written comments on this 

proposed rule must be received by 

September 30, 2011. 

Public Hearings: The EPA intends to 

hold a public hearing around the end of 

August to early September and will 

announce in a separate Federal Register 

notice the date, time, and address of the 

public hearing on this proposed rule. 

ADDRESSES: Submit your comments, 

identified by Docket ID No. EPA–HQ– 

OAR–2007–1145, by one of the 

following methods: 

• http://www.regulations.gov: Follow 

the on-line instructions for submitting 

comments. 

• E-mail: a-and-r-Docket@epa.gov. 

• Fax: 202–566–1741. 

• Mail: Docket No. EPA–HQ–OAR– 

2007–1145, Environmental Protection 

Agency, Mail code 6102T, 1200 

Pennsylvania Ave., NW., Washington, 

DC 20460. Please include a total of two 

copies. 

• Hand Delivery: Docket No. EPA– 

HQ–OAR–2007–1145, Environmental 

Protection Agency, EPA West, Room 

3334, 1301 Constitution Ave., NW., 

Washington, DC. Such deliveries are 

only accepted during the Docket’s 

normal hours of operation, and special 

arrangements should be made for 

deliveries of boxed information. 

Instructions: Direct your comments to 

Docket ID No. EPA–HQ–OAR–2007– 

1145. The EPA’s policy is that all 

comments received will be included in 

the public docket without change and 

may be made available online at http:// 

www.regulations.gov, including any 

personal information provided, unless 

the comment includes information 

claimed to be Confidential Business 

Information (CBI) or other information 

whose disclosure is restricted by statute. 

Do not submit information that you 

consider to be CBI or otherwise 

protected through http:// 

www.regulations.gov or e-mail. The 

http://www.regulations.gov Web site is 

an ‘‘anonymous access’’ system, which 

means EPA will not know your identity 

or contact information unless you 

provide it in the body of your comment. 

If you send an e-mail comment directly 

to EPA without going through http:// 

www.regulations.gov, your e-mail 

address will be automatically captured 

and included as part of the comment 

that is placed in the public docket and 

made available on the Internet. If you 

submit an electronic comment, EPA 

recommends that you include your 

name and other contact information in 

the body of your comment and with any 

disk or CD–ROM you submit. If EPA 

cannot read your comment due to 

technical difficulties and cannot contact 

you for clarification, EPA may not be 

able to consider your comment. 

Electronic files should avoid the use of 

special characters, any form of 

encryption, and be free of any defects or 

viruses. For additional information 

about EPA’s public docket, visit the EPA 

Docket Center homepage at http:// 

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

Docket: All documents in the docket 

are listed in the http:// 

www.regulations.gov index. Although 

listed in the index, some information is 

not publicly available, e.g., CBI or other 

information whose disclosure is 

restricted by statute. Certain other 

material, such as copyrighted material, 

will be publicly available only in hard 

copy. Publicly available docket 

materials are available either 

electronically in http:// 

www.regulations.gov or in hard copy at 

the Air and Radiation Docket and 

Information Center, EPA/DC, EPA West, 

Room 3334, 1301 Constitution Ave., 

NW., Washington, DC. The Public 

Reading Room is open from 8:30 a.m. to 

4:30 p.m., Monday through Friday, 

excluding legal holidays. The telephone 

number for the Public Reading Room is 

(202) 566–1744 and the telephone 

number for the Air and Radiation 

Docket and Information Center is (202) 

566–1742. 

FOR FURTHER INFORMATION CONTACT: Dr. 

Richard Scheffe, Office of Air Quality 

Planning and Standards, U.S. 

Environmental Protection Agency, Mail 

code C304–02, Research Triangle Park, 

NC 27711; telephone: 919–541–4650; 

fax: 919–541–2357; e-mail: 

scheffe.rich@epa.gov. 

SUPPLEMENTARY INFORMATION: 

General Information 

What should I consider as I prepare my 

comments for EPA? 

1. Submitting CBI. Do not submit this 

information to EPA through http:// 

www.regulations.gov or e-mail. Clearly 

mark the part or all of the information 

that you claim to be CBI. For CBI 

information in a disk or CD ROM that 

you mail to EPA, mark the outside of the 

disk or CD ROM as CBI and then 

identify electronically within the disk or 

CD ROM the specific information that is 

claimed as CBI. In addition to one 

complete version of the comment that 

includes information claimed as CBI, a 

copy of the comment that does not 

contain the information claimed as CBI 

must be submitted for inclusion in the 

public docket. Information so marked 

will not be disclosed except in 

accordance with procedures set forth in 

40 CFR part 2. 

2. Tips for Preparing Your Comments. 

When submitting comments, remember 

to: 

• Identify the rulemaking by docket 

number and other identifying 

information (subject heading, Federal 

Register date and page number). 


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• Follow directions—The Agency 

may ask you to respond to specific 

questions or organize comments by 

referencing a Code of Federal 

Regulations (CFR) part or section 

number. 

• Explain why you agree or disagree, 

suggest alternatives, and substitute 

language for your requested changes. 

• Describe any assumptions and 

provide any technical information and/ 

or data that you used. 

• If you estimate potential costs or 

burdens, explain how you arrived at 

your estimate in sufficient detail to 

allow for it to be reproduced. 

• Provide specific examples to 

illustrate your concerns, and suggest 

alternatives. 

• Explain your views as clearly as 

possible. 

• Make sure to submit your 

comments by the comment period 

deadline identified. 

Availability of Related Information 

A number of documents relevant to 

this rulemaking are available on EPA 

web sites. The Integrated Science 

Assessment for Oxides of Nitrogen and 

Sulfur—Ecological Criteria: Final Report 

(ISA) is available on EPAs National 

Center for Environmental Assessment 

Web site. To obtain this document, go 

to http://www.epa.gov/ncea, and click 

on Air Quality then click on Oxides of 

Nitrogen and Sulfur. The Policy 

Assessment (PA), Risk and Exposure 

Assessment (REA), and other related 

technical documents are available on 

EPA’s Office of Air Quality Planning 

and Standards (OAQPS) Technology 

Transfer Network (TTN) web site. The 

PA is available at http://www.epa.gov/ 

ttn/naaqs/standards/no2so2sec/ 

cr_pa.html, and the exposure and risk 

assessments and other related technical 

documents are available at http:// 

www.epa.gov/ttn/naaqs/standards/ 

no2so2sec/cr_rea.html. These and other 

related documents are also available for 

inspection and copying in the EPA 

docket identified above. 

Table of Contents 

The following topics are discussed in 

this preamble: 

I. Background 

A. Legislative Requirements 

B. History of Reviews of NAAQS for 

Nitrogen Oxides and Sulfur Oxides 

1. NAAQS for Oxides of Nitrogen 

2. NAAQS for Oxides of Sulfur 

C. History of Related Assessments and 

Agency Actions 

D. History of the Current Review 

E. Scope of the Current Review 

II. Rationale for Proposed Decision on the 

Adequacy of the Current Secondary 

Standards 

A. Ecological Effects 

1. Effects Associated with Gas-Phase 

Oxides of Nitrogen and Sulfur 

a. Nature of ecosystem responses to gasphase nitrogen and sulfur 

b. Magnitude of ecosystem response to gasphase nitrogen and sulfur 

2. Acidification Effects Associated with 

Deposition of Oxides of Nitrogen and 

Sulfur 

a. Nature of Acidification-related 

Ecosystem Responses 

i. Aquatic Ecosystems 

ii. Terrestrial Ecosystems 

iii. Ecosystem Sensitivity 

b. Magnitude of Acidification-Related 

Ecosystem Responses 

i. Aquatic Acidification 

ii. Terrestrial Acidification 

c. Key Uncertainties Associated With 

Acidification 

i. Aquatic Acidification 

ii. Terrestrial Acidification 

3. Nutrient Enrichment Effects Associated 

With Deposition of Oxides of Nitrogen 

a. Nature of Nutrient Enrichment-Related 

Ecosystem Responses 

i. Aquatic Ecosystems 

ii. Terrestrial Ecosystems 

iii. Ecosystem Eensitivity to Nutrient 

Enrichment 

b. Magnitude of Nutrient EnrichmentRelated Ecosystem Responses 

i. Aquatic Ecosystems 

ii. Terrestrial Ecosystems 

c. Key Uncertainties Associated With 

Nutrient Enrichment 

i. Aquatic Ecosystems 

ii. Terrestrial Ecosystems 

4. Other Ecological Effects 

B. Risk and Exposure Assessment 

1. Overview of Risk and Exposure 

Assessment 

2. Key Findings 

a. Air Quality Analyses 

b. Deposition-Related Aquatic 

Acidification 

c. Deposition-Related Terrestrial 

Acidification 

d. Deposition-Related Aquatic Nutrient 

Enrichment 

e. Deposition-Related Terrestrial Nutrient 

Enrichment 

f. Additional Effects 

3. Conclusions on Effects 

C. Adversity of Effects to Public Welfare 

1. Ecosystem Services 

2. Effects on Ecosystem Services 

a. Aquatic Acidification 

b. Terrestrial Acidification 

c. Nutrient Enrichment 

3. Summary 

D. Adequacy of the Current Standards 

1. Adequacy of the Current Standards for 

Direct Effects 

2. Appropriateness and Adequacy of the 

Current Standards for Deposition-Related 

Effects 

a. Appropriateness 

b. Adequacy of Protection 

i. Aquatic Acidification 

ii. Terrestrial Acidification 

iii. Terrestrial Nutrient Enrichment 

iv. Aquatic Nutrient Enrichment 

v. Other Effects 

3. CASAC Views 

4. Administrator’s Proposed Conclusions 

Concerning Adequacy of Current 

Standard 

III. Rationale for Proposed Decision on 

Alternative Multi-Pollutant Approach to 

Secondary Standards for Aquatic 

Acidification 

A. Ambient Air Indicators 

1. Oxides of Sulfur 

2. Oxides of Nitrogen 

B. Form 

1. Ecological Indicator 

2. Linking ANC to Deposition 

3. Linking Deposition to Ambient Air 

Indicators 

4. Aquatic Acidification Index 

5. Spatial Aggregation 

a. Ecoregion Sensitivity 

b. Representative Ecoregion-Specific 

Factors 

i. Factor F1 

(a) Acid-Sensitive Ecoregions 

(b) Non-Acid Sensitive Ecoregions 

ii. Factor F2 

iii. Factors F3 and F4 

c. Factors in Data-limited Ecoregions 

d. Application to Hawaii, Alaska, and the 

U.S. Territories 

6. Summary of the AAI Form 

C. Averaging Time 

D. Level 

1. Association Between pH Levels and 

Target ANC Levels 

2. ANC Levels Related to Effects on 

Aquatic Ecosystems 

3. Consideration of Episodic Acidity 

4. Consideration of Ecosystem Response 

Time 

5. Prior Examples of Target ANC Levels 

6. Consideration of Public Welfare Benefits 

7. Summary of Alternative Levels 

E. Combined Alternative Levels and Forms 

F. Characterization of Uncertainties 

1. Overview of Uncertainty 

2. Uncertainties Associated with Data Gaps 

3. Uncertainties in Modeled Processes 

4. Applying Knowledge of Uncertainties 

G. CASAC Advice 

H. Administrator’s Proposed Conclusions 

IV. Field Pilot Program and Ambient 

Monitoring 

A. Field Pilot Program 

1. Objectives 

2. Overview of Field Pilot Program 

3. Complementary Measurements 

4. Complementary Areas of Research 

Implementation Challenges 

5. Final Monitoring Plan Development and 

Stakeholder Participation 

B. Evaluation of Monitoring Methods 

1. Potential FRMs for SO2 and p-SO4

2. Potential FRM for NOy

V. Statutory and Executive Order Reviews 

A. Executive Order 12866: Regulatory 

Planning and Review 

B. Paperwork Reduction Act 

C. Regulatory Flexibility Act 

D. Unfunded Mandates Reform Act 

E. Executive Order 13132: Federalism 

F. Executive Order 13175: Consultation 

and Coordination With Indian Tribal 

Governments 

G. Executive Order 13045: Protection of 

Children From Environmental Health 

and Safety Risks 


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1The legislative history of section 109 indicates 

that a primary standard is to be set at ‘‘the 

maximum permissible ambient air level * * * 

which will protect the health of any [sensitive] 

group of the population,’’ and that for this purpose 

‘‘reference should be made to a representative 

sample of persons comprising the sensitive group 

rather than to a single person in such a group.’’ S. 

Rep. No. 91–1196, 91st Cong., 2d Sess. 10 (1970). 

H. Executive Order 13211: Actions That 

Significantly Affect Energy Supply, 

Distribution, or Use 

I. National Technology Transfer and 

Advancement Act 

J. Executive Order 12898: Federal Actions 

To Address Environmental Justice in 

Minority Populations and Low-Income 

Populations References 

I. Background 

A. Legislative Requirements 

Two sections of the Clean Air Act 

(CAA) govern the establishment and 

revision of the NAAQS. Section 108 (42 

U.S.C. section 7408) directs the 

Administrator to identify and list 

certain air pollutants and then to issue 

air quality criteria for those pollutants. 

The Administrator is to list those air 

pollutants that in her ‘‘judgment, cause 

or contribute to air pollution which may 

reasonably be anticipated to endanger 

public health or welfare;’’ ‘‘the presence 

of which in the ambient air results from 

numerous or diverse mobile or 

stationary sources;’’ and ‘‘for which 

* * * [the Administrator] plans to issue 

air quality criteria * * *’’ Air quality 

criteria are intended to ‘‘accurately 

reflect the latest scientific knowledge 

useful in indicating the kind and extent 

of all identifiable effects on public 

health or welfare which may be 

expected from the presence of [a] 

pollutant in the ambient air * * *’’ 42 

U.S.C. 7408(b). Section 109 (42 U.S.C. 

7409) directs the Administrator to 

propose and promulgate ‘‘primary’’ and 

‘‘secondary’’ NAAQS for pollutants for 

which air quality criteria are issued. 

Section 109(b)(1) defines a primary 

standard as one ‘‘the attainment and 

maintenance of which in the judgment 

of the Administrator, based on such 

criteria and allowing an adequate 

margin of safety, are requisite to protect 

the public health.’’ 1 A secondary 

standard, as defined in section 

109(b)(2), must ‘‘specify a level of air 

quality the attainment and maintenance 

of which, in the judgment of the 

Administrator, based on such criteria, is 

requisite to protect the public welfare 

from any known or anticipated adverse 

effects associated with the presence of 

[the] pollutant in the ambient air.’’ 

Welfare effects as defined in section 

302(h) (42 U.S.C. 7602(h)) include, but 

are not limited to, ‘‘effects on soils, 

water, crops, vegetation, man-made 

materials, animals, wildlife, weather, 

visibility and climate, damage to and 

deterioration of property, and hazards to 

transportation, as well as effects on 

economic values and on personal 

comfort and well-being.’’ 

In setting standards that are 

‘‘requisite’’ to protect public health and 

welfare, as provided in section 109(b), 

EPA’s task is to establish standards that 

are neither more nor less stringent than 

necessary for these purposes. In so 

doing, EPA may not consider the costs 

of implementing the standards. See 

generally, Whitman v. American 

Trucking Associations, 531 U.S. 457, 

465–472, 475–76 (2001). Likewise, 

‘‘[a]ttainability and technological 

feasibility are not relevant 

considerations in the promulgation of 

national ambient air quality standards.’’ 

American Petroleum Institute v. Costle, 

665 F. 2d at 1185. Section 109(d)(1) 

requires that ‘‘not later than December 

31, 1980, and at 5-year intervals 

thereafter, the Administrator shall 

complete a thorough review of the 

criteria published under section 108 and 

the national ambient air quality 

standards * * * and shall make such 

revisions in such criteria and standards 

and promulgate such new standards as 

may be appropriate * * * .’’ Section 

109(d)(2) requires that an independent 

scientific review committee ‘‘shall 

complete a review of the criteria * * * 

and the national primary and secondary 

ambient air quality standards * * * and 

shall recommend to the Administrator 

any new * * * standards and revisions 

of existing criteria and standards as may 

be appropriate * * * .’’ Since the early 

1980’s, this independent review 

function has been performed by the 

Clean Air Scientific Advisory 

Committee (CASAC). 

B. History of Reviews of NAAQS for 

Nitrogen Oxides and Sulfur Oxides 

1. NAAQS for Oxides of Nitrogen 

After reviewing the relevant science 

on the public health and welfare effects 

associated with oxides of nitrogen, EPA 

promulgated identical primary and 

secondary NAAQS for NO2 in April 

1971. These standards were set at a level 

of 0.053 parts per million (ppm) as an 

annual average (36 FR 8186). In 1982, 

EPA published Air Quality Criteria 

Document for Oxides of Nitrogen 

(US EPA, 1982), which updated the 

scientific criteria upon which the initial 

standards were based. In February 1984 

EPA proposed to retain these standards 

(49 FR 6866). After taking into account 

public comments, EPA published the 

final decision to retain these standards 

in June 1985 (50 FR 25532). 

The EPA began the most recent 

previous review of the oxides of 

nitrogen secondary standards in 1987. 

In November 1991, EPA released an 

updated draft air quality criteria 

document (AQCD) for CASAC and 

public review and comment (56 FR 

59285), which provided a 

comprehensive assessment of the 

available scientific and technical 

information on health and welfare 

effects associated with NO2 and other 

oxides of nitrogen. The CASAC 

reviewed the draft document at a 

meeting held on July 1, 1993 and 

concluded in a closure letter to the 

Administrator that the document 

‘‘provides a scientifically balanced and 

defensible summary of current 

knowledge of the effects of this 

pollutant and provides an adequate 

basis for EPA to make a decision as to 

the appropriate NAAQS for NO2’’ 

(Wolff, 1993). The AQCD for Oxides of 

Nitrogen was then finalized (US EPA, 

1995a). The EPA’s OAQPS also 

prepared a Staff Paper that summarized 

and integrated the key studies and 

scientific evidence contained in the 

revised AQCD for oxides of nitrogen and 

identified the critical elements to be 

considered in the review of the NO2

NAAQS. The CASAC reviewed two 

drafts of the Staff Paper and concluded 

in a closure letter to the Administrator 

that the document provided a 

‘‘scientifically adequate basis for 

regulatory decisions on nitrogen 

dioxide’’ (Wolff, 1995). 

In October 1995, the Administrator 

announced her proposed decision not to 

revise either the primary or secondary 

NAAQS for NO2 (60 FR 52874; October 

11, 1995). A year later, the 

Administrator made a final 

determination not to revise the NAAQS 

for NO2 after careful evaluation of the 

comments received on the proposal (61 

FR 52852; October 8, 1996). While the 

primary NO2 standard was revised in 

January 2010 by supplementing the 

existing annual standard with the 

establishment of a new 1-hour standard, 

set at a level of 100 ppb (75 FR 6474), 

the secondary NAAQS for NO2 remains 

0.053 ppm (100 micrograms per cubic 

meter [μg/m3] of air), annual arithmetic 

average, calculated as the arithmetic 

mean of the 1-hour NO2 concentrations. 

2. The NAAQS for Oxides of Sulfur 

The EPA promulgated primary and 

secondary NAAQS for SO2 in April 

1971 (36 FR 8186). The secondary 

standards included a standard set at 

0.02 ppm, annual arithmetic mean, and 

a 3-hour average standard set at 0.5 

ppm, not to be exceeded more than once 

per year. These secondary standards 


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were established solely on the basis of 

evidence of adverse effects on 

vegetation. In 1973, revisions made to 

Chapter 5 (‘‘Effects of Sulfur Oxide in 

the Atmosphere on Vegetation’’) of the 

AQCD for Sulfur Oxides (US EPA, 1973) 

indicated that it could not properly be 

concluded that the vegetation injury 

reported resulted from the average SO2

exposure over the growing season, 

rather than from short-term peak 

concentrations. Therefore, EPA 

proposed (38 FR 11355) and then 

finalized (38 FR 25678) a revocation of 

the annual mean secondary standard. At 

that time, EPA was aware that thencurrent concentrations of oxides of 

sulfur in the ambient air had other 

public welfare effects, including effects 

on materials, visibility, soils, and water. 

However, the available data were 

considered insufficient to establish a 

quantitative relationship between 

specific ambient concentrations of 

oxides of sulfur and such public welfare 

effects (38 FR 25679). 

In 1979, EPA announced that it was 

revising the AQCD for oxides of sulfur 

concurrently with that for particulate 

matter (PM) and would produce a 

combined PM and oxides of sulfur 

criteria document. Following its review 

of a draft revised criteria document in 

August 1980, CASAC concluded that 

acid deposition was a topic of extreme 

scientific complexity because of the 

difficulty in establishing firm 

quantitative relationships among (1) 

Emissions of relevant pollutants (e.g., 

SO2 and oxides of nitrogen), (2) 

formation of acidic wet and dry 

deposition products, and (3) effects on 

terrestrial and aquatic ecosystems. The 

CASAC also noted that acid deposition 

involves, at a minimum, several 

different criteria pollutants: Oxides of 

sulfur, oxides of nitrogen, and the fine 

particulate fraction of suspended 

particles. The CASAC felt that any 

document on this subject should 

address both wet and dry deposition, 

since dry deposition was believed to 

account for a substantial portion of the 

total acid deposition problem. 

For these reasons, CASAC 

recommended that a separate, 

comprehensive document on acid 

deposition be prepared prior to any 

consideration of using the NAAQS as a 

regulatory mechanism for the control of 

acid deposition. The CASAC also 

suggested that a discussion of acid 

deposition be included in the AQCDs 

for oxides of nitrogen and PM and 

oxides of sulfur. Following CASAC 

closure on the AQCD for oxides of 

sulfur in December 1981, EPA’s OAQPS 

published a Staff Paper in November 

1982, although the paper did not 

directly assess the issue of acid 

deposition. Instead, EPA subsequently 

prepared the following documents to 

address acid deposition: The Acidic 

Deposition Phenomenon and Its Effects: 

Critical Assessment Review Papers, 

Volumes I and II (US EPA, 1984a, b) and 

The Acidic Deposition Phenomenon 

and Its Effects: Critical Assessment 

Document (US EPA, 1985) (53 FR 

14935–14936). These documents, 

though they were not considered criteria 

documents and did not undergo CASAC 

review, represented the most 

comprehensive summary of scientific 

information relevant to acid deposition 

completed by EPA at that point. 

In April 1988 (53 FR 14926), EPA 

proposed not to revise the existing 

primary and secondary standards for 

SO2. This proposed decision with regard 

to the secondary SO2 NAAQS was due 

to the Administrator’s conclusions that: 

(1) Based upon the then-current 

scientific understanding of the acid 

deposition problem, it would be 

premature and unwise to prescribe any 

regulatory control program at that time; 

and (2) when the fundamental scientific 

uncertainties had been decreased 

through ongoing research efforts, EPA 

would draft and support an appropriate 

set of control measures. Although EPA 

revised the primary SO2 standard in 

June 2010 by establishing a new 1-hour 

standard at a level of 75 ppb and 

revoking the existing 24-hour and 

annual standards (75 FR 35520), no 

further decisions on the secondary SO2

standard have been published. 

C. History of Related Assessments and 

Agency Actions 

In 1980, the Congress created the 

National Acid Precipitation Assessment 

Program (NAPAP) in response to 

growing concern about acidic 

deposition. The NAPAP was given a 

broad 10-year mandate to examine the 

causes and effects of acidic deposition 

and to explore alternative control 

options to alleviate acidic deposition 

and its effects. During the course of the 

program, the NAPAP issued a series of 

publicly available interim reports prior 

to the completion of a final report in 

1990 (NAPAP, 1990). 

In spite of the complexities and 

significant remaining uncertainties 

associated with the acid deposition 

problem, it soon became clear that a 

program to address acid deposition was 

needed. The Clean Air Act Amendments 

of 1990 included numerous separate 

provisions related to the acid deposition 

problem. The primary and most 

important of the provisions, the 

amendments to Title IV of the Act, 

established the Acid Rain Program to 

reduce emissions of SO2 by 10 million 

tons and emissions of nitrogen oxides 

by 2 million tons from 1980 emission 

levels in order to achieve reductions 

over broad geographic regions. In this 

provision, Congress included a 

statement of findings that led them to 

take action, concluding that (1) The 

presence of acid compounds and their 

precursors in the atmosphere and in 

deposition from the atmosphere 

represents a threat to natural resources, 

ecosystems, materials, visibility, and 

public health; (2) the problem of acid 

deposition is of national and 

international significance; and 

(3) current and future generations of 

Americans will be adversely affected by 

delaying measures to remedy the 

problem. 

Second, Congress authorized the 

continuation of the NAPAP in order to 

assure that the research and monitoring 

efforts already undertaken would 

continue to be coordinated and would 

provide the basis for an impartial 

assessment of the effectiveness of the 

Title IV program. 

Third, Congress considered that 

further action might be necessary in the 

long term to address any problems 

remaining after implementation of the 

Title IV program and, reserving 

judgment on the form that action could 

take, included Section 404 of the 1990 

Amendments (Clean Air Act 

Amendments of 1990, Pub. L. 101–549, 

§ 404) requiring EPA to conduct a study 

on the feasibility and effectiveness of an 

acid deposition standard or standards to 

protect ‘‘sensitive and critically 

sensitive aquatic and terrestrial 

resources.’’ At the conclusion of the 

study, EPA was to submit a report to 

Congress. Five years later, EPA 

submitted its report, entitled Acid 

Deposition Standard Feasibility Study: 

Report to Congress (US EPA, 1995b) in 

fulfillment of this requirement. That 

report concluded that establishing acid 

deposition standards for sulfur and 

nitrogen deposition may at some point 

in the future be technically feasible, 

although appropriate deposition loads 

for these acidifying chemicals could not 

be defined with reasonable certainty at 

that time. 

Fourth, the 1990 Amendments also 

added new language to sections of the 

CAA pertaining to the scope and 

application of the secondary NAAQS 

designed to protect the public welfare. 

Specifically, the definition of ‘‘effects on 

welfare’’ in Section 302(h) was 

expanded to state that the welfare 

effects include effects ‘‘* * * whether 

caused by transformation, conversion, 

or combination with other air 

pollutants.’’ 


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2Center for Biological Diversity, et al. v. Johnson, 

No. 05–1814 (D.D.C.) 

In 1999, seven Northeastern states 

cited this amended language in Section 

302(h) in a petition asking EPA to use 

its authority under the NAAQS program 

to promulgate secondary NAAQS for the 

criteria pollutants associated with the 

formation of acid rain. The petition 

stated that this language ‘‘clearly 

references the transformation of 

pollutants resulting in the inevitable 

formation of sulfate and nitrate aerosols 

and/or their ultimate environmental 

impacts as wet and dry deposition, 

clearly signaling Congressional intent 

that the welfare damage occasioned by 

sulfur and nitrogen oxides be addressed 

through the secondary standard 

provisions of Section 109 of the Act.’’ 

The petition further stated that ‘‘recent 

federal studies, including the NAPAP 

Biennial Report to Congress: An 

Integrated Assessment, document the 

continued and increasing damage being 

inflicted by acid deposition to the lakes 

and forests of New York, New England 

and other parts of our nation, 

demonstrating that the Title IV program 

had proven insufficient.’’ The petition 

also listed other adverse welfare effects 

associated with the transformation of 

these criteria pollutants, including 

impaired visibility, eutrophication of 

coastal estuaries, global warming, and 

tropospheric ozone and stratospheric 

ozone depletion. 

In a related matter, the Office of the 

Secretary of the U.S. Department of 

Interior (DOI) requested in 2000 that 

EPA initiate a rulemaking proceeding to 

enhance the air quality in national parks 

and wilderness areas in order to protect 

resources and values that are being 

adversely affected by air pollution. 

Included among the effects of concern 

identified in the request were the 

acidification of streams, surface waters, 

and/or soils; eutrophication of coastal 

waters; visibility impairment; and foliar 

injury from ozone. 

In a Federal Register notice in 2001 

(65 FR 48699), EPA announced receipt 

of these requests and asked for comment 

on the issues raised in them. The EPA 

stated that it would consider any 

relevant comments and information 

submitted, along with the information 

provided by the petitioners and DOI, 

before making any decision concerning 

a response to these requests for 

rulemaking. 

The 2005 NAPAP report states that 

‘‘* * * scientific studies indicate that 

the emission reductions achieved by 

Title IV are not sufficient to allow 

recovery of acid-sensitive ecosystems. 

Estimates from the literature of the 

scope of additional emission reductions 

that are necessary in order to protect 

acid-sensitive ecosystems range from 

approximately 40–80% beyond full 

implementation of Title IV. * * *’’ The 

results of the modeling presented in this 

Report to Congress indicate that broader 

recovery is not predicted without 

additional emission reductions 

(NAPAP, 2005). 

Given the state of the science as 

described in the ISA, REA, and in other 

recent reports, such as the NAPAP 

reports noted above, EPA has decided, 

in the context of evaluating the 

adequacy of the current NO2 and SO2

secondary standards in this review, to 

revisit the question of the 

appropriateness of setting secondary 

NAAQS to address remaining known or 

anticipated adverse public welfare 

effects resulting from the acidic and 

nutrient deposition of these criteria 

pollutants. 

D. History of the Current Review 

The EPA initiated this current review 

in December 2005 with a call for 

information (70 FR 73236) for the 

development of a revised ISA. An 

Integrated Review Plan (IRP) was 

developed to provide the framework 

and schedule as well as the scope of the 

review and to identify policy-relevant 

questions to be addressed in the 

components of the review. The IRP was 

released in 2007 (US EPA, 2007) for 

CASAC and public review. The EPA 

held a workshop in July 2007 on the ISA 

to obtain broad input from the relevant 

scientific communities. This workshop 

helped to inform the preparation of the 

first draft ISA, which was released for 

CASAC and public review in December 

2007; a CASAC meeting was held on 

April 2–3, 2008 to review the first draft 

ISA. A second draft ISA was released for 

CASAC and public review in August 

2008, and was discussed at a CASAC 

meeting held on October 1–2, 2008. The 

final ISA (US EPA, 2008) was released 

in December 2008. 

Based on the science presented in the 

ISA, EPA developed the REA to further 

assess the national impact of the effects 

documented in the ISA. The Draft Scope 

and Methods Plan for Risk/Exposure 

Assessment: Secondary NAAQS Review 

for Oxides of Nitrogen and Oxides of 

Sulfur outlining the scope and design of 

the future REA was prepared for CASAC 

consultation and public review in 

March 2008. A first draft REA was 

presented to CASAC and the public for 

review in August 2008 and a second 

draft was presented for review in June 

2009. The final REA (US EPA, 2009) 

was released in September 2009. A first 

draft PA was released in March 2010 

and reviewed by CASAC on April 1–2, 

2010. In a June 22, 2010 letter to the 

Administrator, CASAC provided advice 

and recommendations to the Agency 

concerning the first draft PA (Russell 

and Samet, 2010a). A second draft PA 

was released to CASAC and the public 

in September 2010 and reviewed by 

CASAC on October 6–7, 2010. The 

CASAC provided advice and 

recommendations to the Agency 

regarding the second draft PA in a 

December 9, 2010 letter (Russell and 

Samet 2010b). The CASAC and public 

comments on the second draft PA were 

considered by EPA staff in developing a 

final PA (US EPA, 2011). CASAC 

requested an additional meeting to 

provide additional advice to the 

Administrator based on the final PA on 

February 15–16, 2011. On January 14, 

2011, EPA released a version of the final 

PA prior to final document production, 

to provide sufficient time for CASAC 

review of the document in advance of 

this meeting. The final PA, 

incorporating final reference checks and 

document formatting, was released in 

February 2011. In a May 17, 2011 letter 

(Russell and Samet, 2011a), CASAC 

offered additional advice and 

recommendations to the Administrator 

with regard to the review of the 

secondary NAAQS for oxides of 

nitrogen and oxides of sulfur. 

In 2005, the Center for Biological 

Diversity and four other plaintiffs filed 

a complaint alleging that EPA had failed 

to complete the current review within 

the period provided by statute.2 The 

schedule for completion of this review 

is governed by a consent decree 

resolving that lawsuit and the 

subsequent extension agreed to by the 

parties. The schedule presented in the 

original consent decree that governs this 

review, entered by the court on 

November 19, 2007, was revised on 

October 22, 2009 to allow for a 17- 

month extension of the schedule. The 

current decree provides that EPA sign 

for publication notices of proposed and 

final rulemaking concerning its review 

of the oxides of nitrogen and oxides of 

sulfur NAAQS no later than July 12, 

2011 and March 20, 2012, respectively. 

This action presents the 

Administrator’s proposed decisions on 

the review of the current secondary 

oxides of nitrogen and oxides of sulfur 

standards. Throughout this preamble a 

number of conclusions, findings, and 

determinations proposed by the 

Administrator are noted. While they 

identify the reasoning that supports this 

proposal, they are only proposals and 

are not intended to be final or 

conclusive in nature. The EPA invites 

general, specific, and/or technical 


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comments on all issues involved with 

this proposal, including all such 

proposed judgments, conclusions, 

findings, and determinations. 

E. Scope of the Current Review 

In conducting this periodic review of 

the secondary NAAQS for oxides of 

nitrogen and oxides of sulfur, as 

discussed in the IRP and REA, EPA 

decided to assess the scientific 

information, associated risks, and 

standards relevant to protecting the 

public welfare from adverse effects 

associated jointly with oxides of 

nitrogen and sulfur. Although EPA has 

historically adopted separate secondary 

standards for oxides of nitrogen and 

oxides of sulfur, EPA is conducting a 

joint review of these standards because 

oxides of nitrogen and sulfur, and their 

associated transformation products are 

linked from an atmospheric chemistry 

perspective, as well as from an 

environmental effects perspective. The 

National Research Council (NRC) has 

recommended that EPA consider 

multiple pollutants, as appropriate, in 

forming the scientific basis for the 

NAAQS (NRC, 2004). As discussed in 

the ISA and REA, there is a strong basis 

for considering these pollutants 

together, building upon EPA’s past 

recognition of the interactions of these 

pollutants and on the growing body of 

scientific information that is now 

available related to these interactions 

and associated ecological effects. 

In defining the scope of this review, 

it must be considered that EPA has set 

secondary standards for two other 

criteria pollutants related to oxides of 

nitrogen and sulfur: Ozone and 

particulate matter (PM). Oxides of 

nitrogen are precursors to the formation 

of ozone in the atmosphere, and under 

certain conditions, can combine with 

atmospheric ammonia to form 

ammonium nitrate, a component of fine 

PM. Oxides of sulfur are precursors to 

the formation of particulate sulfate, 

which is a significant component of fine 

PM in many parts of the U.S. There are 

a number of welfare effects directly 

associated with ozone and fine PM, 

including ozone-related damage to 

vegetation and PM-related visibility 

impairment. Protection against those 

effects is provided by the ozone and fine 

PM secondary standards. This review 

focuses on evaluation of the protection 

provided by secondary standards for 

oxides of nitrogen and sulfur for two 

general types of effects: (1) Direct effects 

on vegetation associated with exposure 

to gaseous oxides of nitrogen and sulfur 

in the ambient air, which are the effects 

that the current NO2 and SO2 secondary 

standards protect against; and (2) effects 

associated with the deposition of oxides 

of nitrogen and sulfur to sensitive 

aquatic and terrestrial ecosystems, 

including deposition in the form of 

particulate nitrate and particulate 

sulfate. 

The ISA focuses on the ecological 

effects associated with deposition of 

ambient oxides of nitrogen and sulfur to 

natural sensitive ecosystems, as 

distinguished from commercially 

managed forests and agricultural lands. 

This focus reflects the fact that the 

majority of the scientific evidence 

regarding acidification and nutrient 

enrichment is based on studies in 

unmanaged ecosystems. Non-managed 

terrestrial ecosystems tend to have a 

higher fraction of nitrogen deposition 

resulting from atmospheric nitrogen (US 

EPA, 2008, section 3.3.2.5). In addition, 

the ISA notes that agricultural and 

commercial forest lands are routinely 

fertilized with amounts of nitrogen that 

exceed air pollutant inputs even in the 

most polluted areas (US EPA, 2008, 

section 3.3.9). This review recognizes 

that the effects of nitrogen deposition in 

managed areas are viewed differently 

from a public welfare perspective than 

are the effects of nitrogen deposition in 

natural, unmanaged ecosystems, largely 

due to the more homogeneous, 

controlled nature of species 

composition and development in 

managed ecosystems and the potential 

for benefits of increased productivity in 

those ecosystems. 

In focusing on natural sensitive 

ecosystems, the PA primarily considers 

the effects of ambient oxides of nitrogen 

and sulfur via deposition on multiple 

ecological receptors. The ISA highlights 

effects including those associated with 

acidification and nitrogen nutrient 

enrichment. With a focus on these 

deposition-related effects, EPA’s 

objective is to develop a framework for 

oxides of nitrogen and sulfur standards 

that incorporates ecologically relevant 

factors and that recognizes the 

interactions between the two pollutants 

as they deposit to sensitive ecosystems. 

The overarching policy objective is to 

develop a secondary standard(s) based 

on the ecological criteria described in 

the ISA and the results of the 

assessments in the REA, and consistent 

with the requirement of the CAA to set 

secondary standards that are requisite to 

protect the public welfare from any 

known or anticipated adverse effects 

associated with the presence of these air 

pollutants in the ambient air. Consistent 

with the CAA, this policy objective 

includes consideration of ‘‘variable 

factors * * * which of themselves or in 

combination with other factors may 

alter the effects on public welfare’’ of 

the criteria air pollutants included in 

this review. 

In addition, we have chosen to focus 

on the effects of ambient oxides of 

nitrogen and sulfur on ecological 

impacts on sensitive aquatic ecosystems 

associated with acidifying deposition of 

nitrogen and sulfur, which is a 

transformation product of ambient 

oxides of nitrogen and sulfur. Based on 

the information in the ISA, the 

assessments presented in the REA, and 

advice from CASAC on earlier drafts of 

this PA (Russell and Samet, 2010a, 

2010b), and as discussed in detail in the 

PA, we have the greatest confidence in 

the causal linkages between oxides of 

nitrogen and sulfur and aquatic 

acidification effects relative to other 

deposition-related effects, including 

terrestrial acidification and aquatic and 

terrestrial nutrient enrichment. 

II. Rationale for Proposed Decision on 

the Adequacy of the Current Secondary 

Standards 

Decisions on retaining or revising the 

current secondary standards for oxides 

of nitrogen and sulfur are largely public 

welfare policy judgments based on the 

Administrator’s informed assessment of 

what constitutes requisite protection 

against adverse effects to public welfare. 

A public welfare policy decision should 

draw upon scientific information and 

analyses about welfare effects, exposure 

and risks, as well as judgments about 

the appropriate response to the range of 

uncertainties that are inherent in the 

scientific evidence and analyses. The 

ultimate determination as to what level 

of damage to ecosystems and the 

services provided by those ecosystems 

is adverse to public welfare is not 

wholly a scientific question, although it 

is informed by scientific studies linking 

ecosystem damage to losses in 

ecosystem services, and information on 

the value of those losses of ecosystem 

services. In reaching such decisions, the 

Administrator seeks to establish 

standards that are neither more nor less 

stringent than necessary for this 

purpose. 

This section presents the rationale for 

the Administrator’s proposed 

conclusions with regard to the adequacy 

of protection and ecological relevance of 

the current secondary standards for 

oxides of nitrogen and sulfur. As 

discussed more fully below, this 

rationale considered the latest scientific 

information on ecological effects 

associated with the presence of oxides 

of nitrogen and oxides of sulfur in the 

ambient air. This rationale also takes 

into account: (1) Staff assessments of the 

most policy-relevant information in the 

ISA and staff analyses of air quality, 


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exposure, and ecological risks, 

presented more fully in the REA and in 

the PA, upon which staff conclusions on 

revisions to the secondary oxides of 

nitrogen and oxides of sulfur standards 

are based; (2) CASAC advice and 

recommendations, as reflected in 

discussions of drafts of the ISA, REA, 

and PA at public meetings, in separate 

written comments, and in CASAC’s 

letters to the Administrator; and (3) 

public comments received during the 

development of these documents, either 

in connection with CASAC meetings or 

separately. 

In developing this rationale, EPA has 

drawn upon an integrative synthesis of 

the entire body of evidence, published 

through early 2008, on ecological effects 

associated with the deposition of oxides 

of nitrogen and oxides of sulfur in the 

ambient air (US EPA, 2008). As 

discussed below in section II.A, this 

body of evidence addresses a broad 

range of ecological endpoints associated 

with ambient levels of oxides of 

nitrogen and oxides of sulfur. In 

considering this evidence, EPA focuses 

on those ecological endpoints, such as 

aquatic acidification, for which the ISA 

judges associations with oxides of 

nitrogen and oxides of sulfur to be 

causal, likely causal, or for which the 

evidence is suggestive that oxides of 

nitrogen and/or sulfur contribute to the 

reported effects. The categories of 

causality determinations have been 

developed in the ISA (US EPA, 2008) 

and are discussed in Section 1.6 of the 

ISA. 

Crucial to this review is the 

development of a form for an 

ecologically relevant standard that 

reflects both the geographically variable 

and deposition-dependent nature of the 

effects. The atmospheric levels of oxides 

of nitrogen and sulfur that afford a 

particular level of ecosystem protection 

are those levels that result in an amount 

of deposition that is less than the 

amount of deposition that a given 

ecosystem can accept without defined 

levels of degradation. 

Drawing from the framework 

developed in the REA, the framework 

we used to structure an ecologically 

meaningful secondary standard in the 

PA and to further develop the indicator, 

form, level, and averaging time of such 

a standard in section III of this proposal 

is depicted below and highlights the 

three key linkages that need to be 

considered in developing an 

ecologically relevant standard. 

The following discussion relies 

heavily on chapters 2 and 3 of the PA. 

The PA 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 PA for consideration in 

addressing public and welfare effects 

associated with the presence of oxides 

of nitrogen and oxides of sulfur in the 

ambient air. Of particular note, chapter 

2 of the PA presents information not 

repeated here that characterizes 

emissions, air quality, deposition and 

water quality. It includes discussions of 

the sources of nitrogen and sulfur in the 

atmosphere as well as current ambient 

air quality monitoring networks and 

models. Additional information in this 

section includes ecological modeling 

and water quality data sources. 

Section II.A presents a discussion of 

the effects associated with oxides of 

nitrogen and sulfur in the ambient air. 

The discussion is organized around the 

types of effects being considered, 

including direct effects of gaseous 

oxides of nitrogen and sulfur, 

deposition-related effects related to 

acidification and nutrient enrichment, 

and other effects such as materials 

damage, climate-related effects and 

mercury methylation. 

Section II.B presents a summary and 

discussion of the risk and exposure 

assessment performed for each of the 

four major effects categories. The REA 

uses case studies representing the broad 

geographic variability of the impacts 

from oxides of nitrogen and sulfur to 

conclude that there are ongoing adverse 

effects in many ecosystems from 

deposition of oxides of nitrogen and 

sulfur and that under current emissions 

scenarios these effects are likely to 

continue. 

Section II.C presents a discussion of 

adversity linking ecological effects to 

measures that can be used to 

characterize the extent to which such 

effects are reasonably considered to be 

adverse to public welfare. This involves 

consideration of how to characterize 

adversity from a public welfare 

perspective. In so doing, consideration 

is given to the concept of ecosystem 

services, the evidence of effects on 

ecosystem services, and how ecosystem 

services can be linked to ecological 

indicators. 

Section II.D presents an assessment of 

the adequacy of the current oxides of 

nitrogen and oxides of sulfur secondary 

standards. Consideration is given to the 

adequacy of protection afforded by the 

current standards for both direct and 

deposition-related effects, as well as to 

the appropriateness of the fundamental 

structure and the basic elements of the 

current standards for providing 

protection from deposition-related 

effects. Considerations as to the extent 

to which deposition-related effects that 

could reasonably be judged to be 

adverse to public welfare are occurring 

under current conditions which are 

allowed by the current standards is also 

considered. Discussion of the structures 

and basic elements of the current NO2

and SO2 secondary standards and 

whether they are adequate to protect 

against such effects is presented. 


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A. Ecological Effects 

This section discusses the known or 

anticipated ecological effects associated 

with oxides of nitrogen and sulfur, 

including the direct effects of gas-phase 

exposure to oxides of nitrogen and 

sulfur (section II.A.1) and effects 

associated with deposition-related 

exposure (sections II.A.2 and 3). Section 

II.A. 2 addresses effects related to 

acidification of aquatic and terrestrial 

ecosystems and section II A.3 addresses 

effects related to nutrient enrichment of 

aquatic and terrestrial ecosystems. 

These sections also address questions 

about the nature and magnitude of 

ecosystem responses to reactive nitrogen 

and sulfur deposition, including 

responses related to acidification, 

nutrient depletion, and, in Section II.A 

4 the mobilization of toxic metals in 

sensitive aquatic and terrestrial 

ecosystems. The uncertainties and 

limitations associated with the evidence 

of such effects are also discussed 

throughout this section. 

1. Effects Associated With Gas-Phase 

Oxides of Nitrogen and Sulfur 

Ecological effects on vegetation as 

discussed in earlier reviews as well as 

the ISA can be attributed to gas-phase 

oxides of nitrogen and sulfur. Acute and 

chronic exposures to gaseous pollutants 

such as SO2, NO2, nitric oxide (NO), 

nitric acid (HNO3) and peroxyacetyl 

nitrite (PAN) are associated with 

negative impacts to vegetation. The 

current secondary NAAQS were set to 

protect against direct damage to 

vegetation by exposure to gas-phase 

oxides of nitrogen and sulfur, such as 

foliar injury, decreased photosynthesis, 

and decreased growth. The following 

summary is a concise overview of the 

known or anticipated effects to 

vegetation caused by gas phase nitrogen 

and sulfur. Most phototoxic effects 

associated with gas phase oxides of 

nitrogen and sulfur occur at levels well 

above ambient concentrations observed 

in the U.S. (US EPA, 2008, section 

3.4.2.4). 

a. Nature of Ecosystem Responses to 

Gas-Phase Nitrogen And Sulfur 

The 2008 ISA found that gas phase 

nitrogen and sulfur are associated with 

direct phytotoxic effects (US EPA, 2008, 

section 4.4). The evidence is sufficient 

to infer a causal relationship between 

exposure to SO2 and injury to vegetation 

(US EPA, 2008, section 4.4.1 and 

3.4.2.1). Acute foliar injury to vegetation 

from SO2 may occur at levels above the 

current secondary standard (3-h average 

of 0.50 ppm). Effects on growth, reduced 

photosynthesis and decreased yield of 

vegetation are also associated with 

increased SO2 exposure concentration 

and time of exposure. 

The evidence is sufficient to infer a 

causal relationship between exposure to 

NO, NO2 and PAN and injury to 

vegetation (US EPA, 2008, section 4.4.2 

and 3.4.2.2). At sufficient 

concentrations, NO, NO2 and PAN can 

decrease photosynthesis and induce 

visible foliar injury to plants. Evidence 

is also sufficient to infer a causal 

relationship between exposure to HNO3

and changes to vegetation (US EPA, 

2008, section 4.4.3 and 3.4.2.3). 

Phytotoxic effects of this pollutant 

include damage to the leaf cuticle in 

vascular plants and disappearance of 

some sensitive lichen species. 

b. Magnitude of Ecosystem Response to 

Gas-Phase Nitrogen And Sulfur 

Vegetation in ecosystems near sources 

of gaseous oxides of nitrogen and sulfur 

or where SO2, NO, NO2, PAN and HNO3

are most concentrated are more likely to 

be impacted by these pollutants. Uptake 

of these pollutants in a plant canopy is 

a complex process involving adsorption 

to surfaces (leaves, stems and soil) and 

absorption into leaves (US EPA, 2008, 

section 3.4.2). The functional 

relationship between ambient 

concentrations of gas phase oxides of 

nitrogen and sulfur and specific plant 

response are impacted by internal 

factors such as rate of stomatal 

conductance and plant detoxification 

mechanisms, and external factors 

including plant water status, light, 

temperature, humidity, and pollutant 

exposure regime (US EPA, 2008, section 

3.4.2). 

Entry of gases into a leaf is dependent 

upon physical and chemical processes 

of gas phase as well as to stomatal 

aperture. The aperture of the stomata is 

controlled largely by the prevailing 

environmental conditions, such as water 

availability, humidity, temperature, and 

light intensity. When the stomata are 

closed, resistance to gas uptake is high 

and the plant has a very low degree of 

susceptibility to injury. Mosses and 

lichens do not have a protective cuticle 

barrier to gaseous pollutants or stomata 

and are generally more sensitive to 

gaseous sulfur and nitrogen than 

vascular plants (US EPA, 2008, section 

3.4.2). 

The appearance of foliar injury can 

vary significantly across species and 

growth conditions affecting stomatal 

conductance in vascular plants (US 

EPA, 2009, section 6.4.1). For example, 

damage to lichens from SO2 exposure 

includes decreased photosynthesis and 

respiration, damage to the algal 

component of the lichen, leakage of 

electrolytes, inhibition of nitrogen 

fixation, decreased potassium (K+) 

absorption, and structural changes. 

The phytotoxic effects of gas phase 

oxides of nitrogen and sulfur are 

dependent on the exposure 

concentration and duration and species 

sensitivity to these pollutants. Effects to 

vegetation associated with oxides of 

nitrogen and sulfur are therefore 

variable across the U.S. and tend to be 

higher near sources of photochemical 

smog. For example, SO2 is considered to 

be the primary factor contributing to the 

death of lichens in many urban and 

industrial areas. 

The ISA states there is very limited 

new research on phytotoxic effects of 

NO, NO2, PAN and HNO3 at 

concentrations currently observed in the 

U.S. with the exception of some lichen 

species (US EPA, 2008, section 4.4). Past 

and current HNO3 concentrations may 

be contributing to the decline in lichen 

species in the Los Angeles basin. Most 

phytotoxic effects associated with gas 

phase oxides of nitrogen and sulfur 

occur at levels well above ambient 

concentrations observed in the U.S. (US 

EPA, 2008, section 3.4.2.4). 

2. Acidification Effects Associated With 

Deposition of Oxides of Nitrogen and 

Sulfur 

Sulfur oxides and nitrogen oxides in 

the atmosphere undergo a complex mix 

of reactions in gaseous, liquid, and solid 

phases to form various acidic 

compounds. These acidic compounds 

are removed from the atmosphere 

through deposition: either wet (e.g., 

rain, snow), fog or cloud, or dry (e.g., 

gases, particles). Deposition of these 

acidic compounds to ecosystems can 

lead to effects on ecosystem structure 

and function. Following deposition, 

these compounds can, in some 

instances, unless retained by soil or 

biota, leach out of the soils in the form 

of sulfate (SO4

2¥) and nitrate (NO3

¥), 

leading to the acidification of surface 

waters. The effects on ecosystems 

depend on the magnitude and rate of 

deposition, as well as a host of 

biogeochemical processes occurring in 

the soils and water bodies (US EPA, 

2009, section 2.1). The chemical forms 

of nitrogen that may contribute to 

acidifying deposition include both 

oxidized and reduced chemical species, 

including reduced forms of nitrogen 

(NHx). 

When sulfur or nitrogen leaches from 

soils to surface waters in the form of 

SO4

2¥ or NO3

¥, an equivalent amount 

of positive cations, or countercharge, is 

also transported. This maintains 

electroneutrality. If the countercharge is 

provided by base cations, such as 


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calcium (Ca2+), magnesium (Mg2+), 

sodium (Na+), or K+, rather than 

hydrogen (H+) and dissolved inorganic 

aluminum, the acidity of the soil water 

is neutralized, but the base saturation of 

the soil decreases. Continued SO4

2

or NO3

¥ leaching can deplete the 

available base cation pool in soil. As the 

base cations are removed, continued 

deposition and leaching of SO4

2¥ and/ 

or NO3

¥ (with H+ and Al3+) leads to 

acidification of soil water, and by 

connection, surface water. Introduction 

of strong acid anions such as sulfate and 

nitrate to an already acidic soil, whether 

naturally or due to anthropogenic 

activities, can lead to instantaneous 

acidification of waterbodies through 

direct runoff without any significant 

change in base cation saturation. The 

ability of a watershed to neutralize 

acidic deposition is determined by a 

variety of biogeophysical factors 

including weathering rates, bedrock 

composition, vegetation and microbial 

processes, physical and chemical 

characteristics of soils and hydrologic 

flowpaths (US EPA, 2009, section 2.1). 

Some of these factors such as vegetation 

and soil depth are highly variable over 

small spatial scales such as meters, but 

can be aggregated to evaluate patterns 

over larger spatial scales. Acidifying 

deposition of oxides of nitrogen and 

sulfur and the chemical and biological 

responses associated with these inputs 

vary temporally. Chronic or long-term 

deposition processes in the time scale of 

years to decades result in increases in 

inputs of nitrogen and sulfur to 

ecosystems and the associated 

ecological effects. Episodic or short term 

(i.e., hours or days) deposition refers to 

events in which the level of the acid 

neutralizing capacity (ANC) of a lake or 

stream is temporarily lowered. In 

aquatic ecosystems, short-term (i.e., 

hours or days) episodic changes in 

water chemistry can have significant 

biological effects. Episodic acidification 

refers to conditions during precipitation 

or snowmelt events when 

proportionately more drainage water is 

routed through upper soil horizons that 

tend to provide less acid neutralizing 

than is passing through deeper soil 

horizons (US EPA, 2009, section 4.2). In 

addition, the accumulated sulfate and 

nitrate in snow packs can provide a 

surge of acidic inputs. Some streams 

and lakes may have chronic or base flow 

chemistry that is suitable for aquatic 

biota, but may be subject to occasional 

acidic episodes with deleterious 

consequences to sensitive biota. 

The following summary is a concise 

overview of the known or anticipated 

effects caused by acidification to 

ecosystems within the U.S. 

Acidification affects both terrestrial and 

freshwater aquatic ecosystems. 

a. Nature of Acidification-Related 

Ecosystem Responses 

The ISA concluded that deposition of 

oxides of nitrogen and sulfur and NHx

leads to the varying degrees of 

acidification of ecosystems (US EPA, 

2008). In the process of acidification, 

biogeochemical components of 

terrestrial and freshwater aquatic 

ecosystems are altered in a way that 

leads to effects on biological organisms. 

Deposition to terrestrial ecosystems 

often moves through the soil and 

eventually leaches into adjacent water 

bodies. 

i. Aquatic Ecosystems 

The scientific evidence is sufficient to 

infer a causal relationship between 

acidifying deposition and effects on 

biogeochemistry and biota in aquatic 

ecosystems (US EPA, 2008, section 

4.2.2). The strongest evidence comes 

from studies of surface water chemistry 

in which acidic deposition is observed 

to alter sulfate and nitrate 

concentrations in surface waters, the 

sum of base cations, ANC, dissolved 

inorganic aluminum and pH (US EPA, 

2008, section 3.2.3.2). The ANC is a key 

indicator of acidification with relevance 

to both terrestrial and aquatic 

ecosystems. The ANC is useful because 

it integrates the overall acid-base status 

of a lake or stream and reflects how 

aquatic ecosystems respond to acidic 

deposition over time. There is also a 

relationship between ANC and the 

surface water constituents that directly 

contribute to or ameliorate acidityrelated stress, in particular, 

concentrations of hydrogen ion (as pH), 

Ca2+ and aluminum (Al). Moreover, low 

pH surface waters leach aluminum from 

soils, which is quite lethal to fish and 

other aquatic organisms. In aquatic 

systems, there is a direct relationship 

between ANC and fish and phytozooplankton diversity and abundance. 

Low ANC coincides with effects on 

aquatic systems (e.g., individual species 

fitness loss or death, reduced species 

richness, altered community structure). 

At the community level, species 

richness is positively correlated with pH 

and ANC because energy cost in 

maintaining physiological homeostasis, 

growth, and reproduction is high at low 

ANC levels. For example, there is a 

logistic relationship between fish 

species richness and ANC class for 

Adirondack Case Study Area lakes that 

indicates the probability of occurrence 

of an organism for a given value of ANC. 

Biota are generally not harmed when 

ANC values are >100 microequivalents 

per liter (μeq/L). The number of fish 

species also peaks at ANC values >100 

μeq/L. Below 100 μeq/L ANC, fish 

fitness and community diversity begin 

to decline (US EPA, section 4.2). 

Specifically at ANC levels between 100 

and 50 μeq/L, the fitness of sensitive 

species (e.g., brook trout, zooplankton) 

begins to decline. When ANC 

concentrations are <50 μeq/L, they are 

generally associated with death or loss 

of fitness of biota that are sensitive to 

acidification. 

Consistent and coherent 

documentation from multiple studies on 

various species from all major trophic 

levels of aquatic systems shows that 

geochemical alteration caused by 

acidification can result in the loss of 

acid-sensitive biological species (US 

EPA, 2008, section 3.2.3.3). This is most 

often discussed with relation to pH. For 

example, in the Adirondacks, of the 53 

fish species recorded in Adirondack 

lakes about half (26 species) were absent 

from lakes with pH below 6.0. 

Biological effects are linked to changes 

in water chemistry including decreases 

in ANC and pH and increases in 

inorganic Al concentration. The direct 

biological effects are caused by lowered 

pH which leads to increased inorganic 

Al concentrations (US EPA, 2011, 

Figures 3–1 and 3–2). While ANC level 

does not cause direct biological harm it 

is a good overall indicator of the risk of 

acidification (US EPA, 2011, section 

3.1.3). 

There are clear associations between 

ANC, pH and aquatic species mortality 

and health which are summarized in 

section 3.1.1 of the PA. Significant harm 

to sensitive aquatic species has been 

observed at pH levels below 6. Normal 

stream pH levels with little to no 

toxicity range from 6 to 7 (MacAvoy et 

al, 1995). Baker et al (1990) observed 

that ‘‘lakes with pH less than 

approximately 6.0 contain significantly 

fewer species than lakes with pH levels 

above 6.0.’’ As noted in Chapter 3, 

typically at pH <4.5 and an ANC <0 

μeq/L, complete to near-complete loss of 

many taxa of organisms occur, including 

fish and aquatic insect populations, 

whereas other taxa are reduced to only 

acidophilic species. Acid Neutralizing 

Capacity is a measure of how much acid 

can be neutralized in a specific surface 

water system. An ANC value of 0 or 

below means that surface waters have 

no ability to neutralize any additional 

acid inputs. 

Additional evidence can help refine 

the understanding of effects occurring at 

pH levels between 4.5 and 6. When pH 

levels are below 5.6, relatively lower 

trout survival rates were observed in the 


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Shenandoah National Park. In field 

observations, when pH levels dropped 

to 5, mortality rates went to 100 percent 

(Bulger et al, 2000). At pH levels ranging 

from 5.4 to 5.8, cumulative mortality 

continues to increase. Several studies 

have shown that trout exposed to water 

with varying pH levels and fish larvae 

showed increasing mortality as pH 

levels decrease. In one study almost 100 

percent mortality was observed at a pH 

of 4.5 compared to almost 100 percent 

survival at a pH of 6.5. Intermediate pH 

values (6.0, 5.5) in all cases showed 

reduced survival compared with the 

control (6.5), but not by statistically 

significant amounts (US EPA, 2008, 

section 3.2.3.3). 

One important indicator of acid stress 

is increased fish mortality. The response 

of fish to pH is not uniform across 

species. A number of synoptic surveys 

indicated loss of species diversity and 

absence of several fish species in the pH 

range of 5.0 to 5.5. If pH is lower, there 

is a greater likelihood that more fish 

species could be lost without 

replacement, resulting in decreased 

richness and diversity. In general, 

populations of salmonids are not found 

at pH levels less than 5.0, and 

smallmouth bass (Micropterus 

dolomieu) populations are usually not 

found at pH values less than about 5.2 

to 5.5. From Table 3–1, only one study 

showed significant mortality effects 

above a pH of 6, while a number of 

studies showed significant mortality 

when pH levels are at or below 5.5. 

The highest pH level for any of the 

studies reported in the ISA is 6.0, 

suggesting that pH above 6.0 is 

protective against mortality effects for 

most species. Most thresholds are in the 

range of pH of 5.0 to 6.0, which suggests 

that a target pH should be no lower than 

5.0. Protection against mortality in some 

recreationally important species such as 

lake trout (pH threshold of 5.6) and 

crappie (pH threshold of 5.5), combined 

with the evidence of effects on larval 

and embryo survival suggests that pH 

levels greater than 5.5 should be 

targeted to provide protection against 

mortality effects throughout the life 

stages of fish. 

Non-lethal effects have been observed 

at pH levels as high as 6. A study in the 

Shenandoah National Park found that 

the condition factor, a measure of fish 

health expressed as fish weight/length 

multiplied by a scaling constant, is 

positively correlated with stream pH 

levels, and that the condition factor is 

reduced in streams with a pH of 6.0 (US 

EPA, 2008, section 3.2.3.3). 

Biodiversity is another indicator of 

aquatic ecosystem health. A key study 

in the Adirondacks found that lakes 

with a pH of 6.0 had only half the 

potential species of fish (27 of 53 

potential species). There is often a 

positive relationship between pH and 

number of fish species, at least for pH 

values between about 5.0 and 6.5, or 

ANC values between about 0 to 100 μeq/ 

L. Such observed relationships are 

complicated, however, by the tendency 

for smaller lakes and streams, having 

smaller watersheds, to also support 

fewer fish species, irrespective of acidbase chemistry. This pattern may be due 

to a decrease in the number of available 

niches as stream or lake size decreases. 

Nevertheless, fish species richness is 

relatively easily determined and is one 

of the most useful indicators of 

biological effects of surface water 

acidification. 

Changes in stream water pH and ANC 

also contribute to declines in taxonomic 

richness of zooplankton, and 

macroinvertebrates which are often 

sources of food for fish, birds and other 

animal species in various ecosystems. 

These fish may also serve as a source of 

food and recreation for humans. 

Acidification of ecosystems has been 

shown to disrupt food web dynamics 

causing alteration to the diet, breeding 

distribution, and reproduction of certain 

species of birds (US EPA, 2008, section 

4.2.2.2. and Table 3–9). For example, 

breeding distribution of the common 

goldeneye (Bucephala clangula), an 

insectivorous duck, may be affected by 

changes in acidifying deposition. 

Similarly, decreases in prey diversity 

and quantity have been observed to 

create feeding problems for nesting pairs 

of loons on low-pH lakes in the 

Adirondacks. 

ii. Terrestrial Ecosystems 

In terrestrial ecosystems, the evidence 

is sufficient to infer a causal 

relationship between acidifying 

deposition and changes in 

biogeochemistry (US EPA, 2008, section 

4.2.1.1). The strongest evidence comes 

from studies of forested ecosystems, 

with supportive information on other 

plant taxa, including shrubs and lichens 

(US EPA, 2008, section 3.2.2.1.). Three 

useful indicators of chemical changes 

and acidification effects on terrestrial 

ecosystems, showing consistency and 

coherence among multiple studies are: 

soil base saturation, Al concentrations 

in soil water, and soil carbon to nitrogen 

(C:N) ratio (US EPA, 2008, section 

3.2.2.2). 

As discussed in the ISA and REA, in 

soils with base saturation less than 

about 15 to 20 percent, exchange 

chemistry is dominated by Al. Under 

these conditions, responses to inputs of 

sulfuric acid and HNO3 largely involve 

the release and mobilization of 

dissolved inorganic Al. The effect can 

be neutralized by weathering from 

geologic parent material or base cation 

exchange. The Ca2+ and Al 

concentrations in soil water are strongly 

influenced by soil acidification and both 

have been shown to have quantitative 

links to tree health, including Al 

interference with Ca2+ uptake and Al 

toxicity to roots. Effects of nitrification 

and associated acidification and cation 

leaching have been consistently shown 

to occur only in soils with a C:N ratio 

below about 20 to 25. 

Soil acidification caused by acidic 

deposition has been shown to cause 

decreased growth and increased 

susceptibility to disease and injury in 

sensitive tree species. Red spruce (Picea 

rubens) dieback or decline has been 

observed across high elevation areas in 

the Adirondack, Green and White 

mountains. The frequency of freezing 

injury to red spruce needles has 

increased over the past 40 years, a 

period that coincided with increased 

emissions of sulfur and nitrogen oxides 

and increased acidifying deposition. 

Acidifying deposition can contribute to 

dieback in sugar maple (Acer 

saccharum) through depletion of cations 

from soil with low levels of available 

Ca. Grasslands are likely less sensitive 

to acidification than forests due to 

grassland soils being generally rich in 

base cations. 

iii. Ecosystem Sensitivity 

The intersection between current 

deposition loading, historic loading and 

sensitivity defines the ecological 

vulnerability to the effects of 

acidification. Freshwater aquatic and 

some terrestrial ecosystems, notably 

forests, are the ecosystem types which 

are most sensitive to acidification. The 

ISA reports that the principal factor 

governing the sensitivity of terrestrial 

and aquatic ecosystems to acidification 

from sulfur and nitrogen deposition is 

geology (particularly surficial geology). 

Geologic formations having low base 

cation supply generally underlie the 

watersheds of acid-sensitive lakes and 

streams. Other factors that contribute to 

the sensitivity of soils and surface 

waters to acidifying deposition include 

topography, soil chemistry, land use, 

and hydrologic flowpaths. Episodic and 

chronic acidification tends to occur in 

areas that have base-poor bedrock, high 

relief, and shallow soils (US EPA, 2008, 

section 3.2.4.1). 

b. Magnitude of Acidification-Related 

Ecosystem Responses 

Terrestrial and aquatic ecosystems 

differ in their response to acidifying 


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46094 Federal Register / Vol. 76, No. 147 / Monday, August 1, 2011 / Proposed Rules 

deposition. Therefore the magnitude of 

ecosystem response is described 

separately for aquatic and terrestrial 

ecosystems in the following sections. 

The magnitude of response refers to 

both the severity of effects and the 

spatial extent of the U.S. which is 

affected. 

i. Aquatic Acidification 

Freshwater ecosystem surveys and 

monitoring in the eastern U.S. have 

been conducted by many programs 

since the mid-1980s, including EPA’s 

Environmental Monitoring and 

Assessment Program (EMAP), National 

Surface Water Survey (NSWS), 

Temporally Integrated Monitoring of 

Ecosystems (TIME), and Long-term 

Monitoring (LTM) programs. Based on 

analyses of surface water data from 

these programs, New England, the 

Adirondack Mountains, the 

Appalachian Mountains (northern 

Appalachian Plateau and Ridge/Blue 

Ridge region) and the Upper Midwest 

contain the most sensitive lakes and 

streams (i.e., ANC less than about 50 

μeq/L). Portions of northern Florida also 

contain many acidic and low-ANC lakes 

and streams, although the role of 

acidifying deposition in this region is 

less clear. The western U.S. contains 

many of the surface waters most 

sensitive to potential acidification 

effects, but with the exception of the Los 

Angeles Basin and surrounding areas, 

the levels of acidifying deposition are 

low in most areas. Therefore, 

acidification of surface waters by acidic 

deposition is not as prevalent in the 

western U.S., and the extent of chronic 

surface water acidification that has 

occurred in that region to date has likely 

been very limited relative to the Eastern 

U.S. (US EPA, 2008, section 3.2.4.2 and 

US EPA, 2009, section 4.2.2). 

There are a number of species 

including fish, aquatic insects, other 

invertebrates and algae that are sensitive 

to acidification and cannot survive, 

compete or reproduce in acidic waters 

(US EPA, 2008, section 3.2.3.3). 

Decreases in ANC and pH have been 

shown to contribute to declines in 

species richness and declines in 

abundance of zooplankton, 

macroinvertebrates, and fish. Reduced 

growth rates have been attributed to 

acid stress in a number of fish species 

including Atlantic salmon (Salmo 

salar), Chinook salmon (Oncorhynchus 

tshawytscha), lake trout (Salvelinus 

namaycush), rainbow trout 

(Oncorhynchis mykiss), brook trout 

(Salvelinus Fontinalis), and brown trout 

(Salmo trutta). In response to small to 

moderate changes in acidity, acidsensitive species are often replaced by 

other more acid-tolerant species, 

resulting in changes in community 

composition and richness. The effects of 

acidification are continuous, with more 

species being affected at higher degrees 

of acidification. At a point, typically a 

pH <4.5 and an ANC <0 μeq/L, complete 

to near-complete loss of many taxa of 

organisms occur, including fish and 

aquatic insect populations, whereas 

other taxa are reduced to only 

acidophilic species. These changes in 

taxa composition are associated with the 

high energy cost in maintaining 

physiological homeostasis, growth, and 

reproduction at low ANC levels (US 

EPA, 2008, section 3.2.3.3). Decreases in 

species richness related to acidification 

have been observed in the Adirondack 

Mountains and Catskill Mountains of 

New York, New England and 

Pennsylvania, and Virginia. From the 

sensitive areas identified by the ISA, 

further ‘‘case study’’ analyses on aquatic 

ecosystems in the Adirondack 

Mountains and Shenandoah National 

Park were conducted to better 

characterize ecological risk associated 

with acidification (US EPA, 2009, 

section 4). 

The ANC is the most widely used 

indicator of acid sensitivity and has 

been found in various studies to be the 

best single indicator of the biological 

response and health of aquatic 

communities in acid-sensitive systems 

(Lien et al., 1992; Sullivan et al., 2006; 

US EPA, 2008). In the REA, surface 

water trends in SO4

2¥ and NO3

¥

concentrations and ANC levels were 

analyzed to affirm the understanding 

that reductions in deposition could 

influence the risk of acidification. The 

ANC values have been categorized 

according to their effects on biota, as 

shown in the table below. Monitoring 

data from TIME/LTM and EMAP 

programs were assessed for the years 

1990 to 2006, and past, present and 

future water quality levels were 

estimated by both steady-state and 

dynamic biogeochemical models. 

TABLE II–1—ECOLOGICAL EFFECTS ASSOCIATED WITH ALTERNATIVE LEVELS OF ACID NEUTRALIZING CAPACITY (ANC) 

[Source: USEPA, Acid Rain Program] 

Category Label ANC Levels and Expected Ecological Effects 

Acute Concern ................................ <0 μeq/L ........................................ Complete loss of fish populations is expected. Planktonic communities have extremely low diversity and are dominated by acidophilic taxa. The numbers of individuals in plankton species that are 

present are greatly reduced. 

Severe Concern .............................. 0–20 μeq/L .................................... Highly sensitive to episodic acidification. During episodes of high 

acidifying deposition, brook trout populations may experience lethal 

effects. The diversity and distribution of zooplankton communities 

decline sharply. 

Elevated Concern ............................ 20–50 μeq/L .................................. Fish species richness is greatly reduced (i.e., more than half of expected species can be missing). On average, brook trout populations experience sublethal effects, including loss of health, ability 

to reproduce, and fitness. Diversity and distribution of zooplankton 

communities decline. 

Moderate Concern .......................... 50–100 μeq/L ................................ Fish species richness begins to decline (i.e., sensitive species are 

lost from lakes). Brook trout populations are sensitive and variable, 

with possible sublethal effects. Diversity and distribution of 

zooplankton communities also begin to decline as species that are 

sensitive to acidifying deposition are affected. 

Low Concern ................................... >100 μeq/L .................................... Fish species richness may be unaffected. Reproducing brook trout 

populations are expected where habitat is suitable. Zooplankton 

communities are unaffected and exhibit expected diversity and distribution. 


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Studies on fish species richness in the 

Adirondacks Case Study Area 

demonstrated the effect of acidification. 

Of the 53 fish species recorded in 

Adirondack Case Study Area lakes, only 

27 species were found in lakes with a 

pH <6.0. The 26 species missing from 

lakes with a pH <6.0 include important 

recreational species, such as Atlantic 

salmon, tiger trout (Salmo trutta X 

Salvelinus fontinalis), redbreast sunfish 

(Lepomis auritus), bluegill (Lepomis 

macrochirus), tiger musky (Esox 

masquinongy X lucius), walleye (Sander 

vitreus), alewife (Alosa 

pseudoharengus), and kokanee 

(Oncorhynchus nerka), as well as 

ecologically important minnows that are 

commonly consumed by sport fish. A 

survey of 1,469 lakes in the late 1980s 

found 346 lakes to be devoid of fish. 

Among lakes with fish, there was a 

relationship between the number of fish 

species and lake pH, ranging from about 

one species per lake for lakes having a 

pH <4.5 to about six species per lake for 

lakes having a pH >6.5. In the 

Adirondacks, a positive relationship 

exists between the pH and ANC in lakes 

and the number of fish species present 

in those lakes (US EPA, 2008, section 

3.2.3.4). 

Since the mid-1990s, streams in the 

Shenandoah Case Study Area have 

shown slight declines in NO3

¥ and 

SO4

2¥ concentrations in surface waters. 

The 2006 concentrations are still above 

pre-acidification (1860) conditions. 

Model of Acidification of Groundwater 

in Catchments (MAGIC) modeling 

predicts surface water concentrations of 

NO3

¥ and SO4

2¥ are 10- and 32-fold 

higher, respectively, in 2006 than in 

1860. The estimated average ANC across 

60 streams in the Shenandoah Case 

Study Area is 57.9 μeq/L (± 4.5 μeq/L). 

Fifty-five percent of all monitored 

streams in the Shenandoah Case Study 

Area have a current risk of Elevated, 

Severe, or Acute. Of the 55 percent, 18 

percent are chronically acidic today (US 

EPA, 2009, section 4.2.4.3). 

Based on a deposition scenario for 

this study area that maintains current 

emission levels from 2020 to 2050, the 

simulation forecast indicates that a large 

number of streams would still have 

Elevated to Acute problems with acidity 

in 2050. 

Biological effects of increased 

acidification documented in the 

Shenandoah Case Study Area include a 

decrease in the condition factor in 

blacknose dace and a decrease in fish 

biodiversity associated with decreasing 

stream ANC. On average, the fish 

species richness is lower by one fish 

species for every 21 μeq/L decrease in 

ANC in Shenandoah National Park 

streams (US EPA, 2008, section 3.2.3.4). 

ii. Terrestrial Acidification 

The ISA identified a variety of 

indicators that can be used to measure 

the effects of acidification in soils. Most 

effects of terrestrial acidification are 

observed in sensitive forest ecosystem 

in the U.S. Tree health has been linked 

to the availability of base cations (BC) in 

soil (such as Ca2+, Mg2+ and K+), as well 

as soil aluminum (Al) content. Tree 

species show a range of sensitivities to 

Ca/Al and BC/Al soil molar ratios, 

therefore these are good chemical 

indicators because they directly relate to 

the biological effects. Critical BC/Al 

molar ratios for a large variety of tree 

species ranged from 0.2 to 0.8. This 

range is similar to critical ratios of Ca/ 

Al. Plant toxicity or nutrient antagonism 

was reported to occur at Ca/Al molar 

ratios ranging from 0.2 to 2.5 (US EPA, 

2009). 

There has been no systematic national 

survey of terrestrial ecosystems to 

determine the extent and distribution of 

terrestrial ecosystem sensitivity to the 

effects of acidifying deposition. 

However, one preliminary national 

evaluation estimated that ∼15 percent of 

forest ecosystems in the U.S. exceed the 

estimated critical load based on soil 

ANC leaching for sulfur and nitrogen 

deposition by >250 eq/ha/yr (McNulty 

et al., 2007). Forests of the Adirondack 

Mountains of New York, Green 

Mountains of Vermont, White 

Mountains of New Hampshire, the 

Allegheny Plateau of Pennsylvania and 

high-elevation forest ecosystems in the 

southern Appalachians are the regions 

most sensitive to terrestrial acidification 

effects from acidifying deposition (US 

EPA, 2008, section 3.2.4.2). While 

studies show some recovery of surface 

waters, there are widespread 

measurements of ongoing depletion of 

exchangeable base cations in forest soils 

in the northeastern U.S. despite recent 

decreases in acidifying deposition, 

indicating a slow recovery time. 

In the REA, a critical load analysis 

was performed for sugar maple and red 

spruce forests in the eastern U.S. by 

using BC/Al ratio in acidified forest 

soils as an indicator to assess the impact 

of nitrogen and sulfur deposition on tree 

health. These are the two most 

commonly studied tree species in North 

America for effects of acidification. At a 

BC/Al ratio of 1.2, red spruce growth 

can be decreased by 20 percent. Sugar 

maple growth can be decreased by 20 

percent at a BC/Al ratio of 0.6 (US EPA, 

2009, section 4.4). The REA analysis 

determined the health of at least a 

portion of the sugar maple and red 

spruce growing in the U.S. may have 

been compromised with acidifying total 

nitrogen and sulfur deposition. 

Specifically, total nitrogen and sulfur 

deposition levels exceeded three 

selected critical loads for tree growth in 

3 percent to 75 percent of all sugar 

maple plots across 24 states—that is, it 

exceeded the highest (least stringent) of 

the three critical loads in 3 percent of 

plots, and the lowest (most stringent) in 

75 percent of plots. For red spruce, total 

nitrogen and sulfur deposition levels 

exceeded three selected critical loads in 

3 percent to 36 percent of all red spruce 

plots across eight states (US EPA, 2009, 

section 4.4). 

c. Key Uncertainties Associated With 

Acidification 

There are different levels of 

uncertainty associated with 

relationships between deposition, 

ecological effects and ecological 

indicators. In Chapter 7 of the REA, the 

case study analyses associated with 

each targeted effect area were 

synthesized by identifying the strengths, 

limitations, and uncertainties associated 

with the available data, modeling 

approach, and relationship between the 

selected ecological indicator and 

atmospheric deposition as described by 

the ecological effect function (US EPA, 

2009, Figure 1–1). A further discussion 

of uncertainty in aquatic and terrestrial 

ecosystems is presented below. The key 

uncertainties were characterized as 

follows to evaluate the strength of the 

scientific basis for setting a national 

standard to protect against a given effect 

(US EPA, 2009, section 7): 

(1) Data Availability: High, medium or 

low quality. This criterion is based on 

the availability and robustness of data 

sets, monitoring networks, availability 

of data that allows for extrapolation to 

larger assessment areas and input 

parameters for modeling and developing 

the ecological effect function. The 

scientific basis for the ecological 

indicator selected is also incorporated 

into this criterion. 

(2) Modeling Approach: High, fairly 

high, intermediate, or low confidence. 

This value is based on the strengths and 

limitations of the models used in the 

analysis and how accepted they are by 

the scientific community for their 

application in this analysis. 

(3) Ecological Effect Function: High, 

fairly high, intermediate or low 

confidence. This ranking is based on 

how well the ecological effect function 

describes the relationship between 

atmospheric deposition and the 

ecological indicator of an effect. 


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i. Aquatic Acidification 

The REA concludes that the available 

data are robust and considered high 

quality. There is high confidence about 

the use of these data and their value for 

extrapolating to a larger regional 

population of lakes. The EPA TIME/ 

LTM network represents a source of 

long-term, representative sampling. Data 

on sulfate concentrations, nitrate 

concentrations and ANC from 1990 to 

2006 used for this analysis as well as 

EPA EMAP and Regional Environmental 

Monitoring and Assessment Program 

(REMAP) surveys, provide considerable 

data on surface water trends. 

There is fairly high confidence 

associated with modeling and input 

parameters. Uncertainties in water 

quality estimates (i.e., ANC) from 

MAGIC were derived from multiple site 

calibrations. Pre-acidification refers to 

retrospective modeling to estimate water 

quality conditions before man-made 

contributions of acidifying inputs. The 

models are evaluated under current 

conditions to determine how well they 

replicate observed ANC values. The 95 

percent confidence interval for preacidification of lakes was an average of 

15 μeq/L difference in ANC 

concentrations, or 10 percent, and 8 

μeq/L, or 5 percent, for streams (US 

EPA, 2009, section 7.1.2). The use of the 

critical load model to estimate aquatic 

critical loads is limited by the 

uncertainties associated with runoff and 

surface water measurements and in 

estimating the catchment supply of base 

cations from the weathering of bedrock 

and soils (McNulty et al., 2007). 

ii. Terrestrial Acidification 

The available data used to quantify 

the targeted effect of terrestrial 

acidification are robust and considered 

high quality. The U.S. Forest ServiceKane Experimental Forest and 

significant amounts of research work in 

the Allegheny Plateau have produced 

extensive, peer-reviewed data sets. 

Sugar maple and red spruce were the 

focus of the REA since they are 

demonstrated to be negatively affected 

by soil available Ca2+ depletion and 

high concentrations of available Al, and 

occur in areas that receive high 

acidifying deposition. There is high 

confidence about the use of the REA 

terrestrial acidification data and their 

value for extrapolating to a larger 

regional population of forests. 

There is high confidence associated 

with the models, input parameters, and 

assessment of uncertainty used in the 

case study for terrestrial acidification. 

The Simple Mass Balance (SMB) model, 

a commonly used and widely applied 

approach for estimating critical loads, 

was used in the REA analysis (US EPA, 

2008, section 7.2.2). There is fairly high 

confidence associated with the 

ecological effect function developed for 

terrestrial acidification (US EPA, 2009, 

section 7.2.3). 

3. Nutrient Enrichment Effects 

Associated With Deposition of Oxides of 

Nitrogen 

The following summary is a concise 

overview of the known or anticipated 

effects caused by nitrogen nutrient 

enrichment to ecosystems within the 

United States. Nutrient-enrichment 

affects terrestrial, freshwater and 

estuarine ecosystems. Nitrogen 

deposition is a major source of 

anthropogenic nitrogen. For many 

terrestrial and freshwater ecosystems 

other sources of nitrogen including 

fertilizer and waste treatment are greater 

than deposition. Nitrogen deposition 

often contributes to nitrogen-enrichment 

effects in estuaries, but does not drive 

the effects since other sources of 

nitrogen greatly exceed nitrogen 

deposition. Both oxides of nitrogen and 

NHX contribute to nitrogen deposition. 

For the most part, nitrogen effects on 

ecosystems do not depend on whether 

the nitrogen is in oxidized or reduced 

form. Thus, this summary focuses on 

the effects of nitrogen deposition in 

total. 

a. Nature of Nutrient EnrichmentRelated Ecosystem Responses 

The ISA found that deposition of 

nitrogen, including oxides of nitrogen 

and NHX, leads to the nitrogen 

enrichment of ecosystems (US EPA 

2008). In the process of nitrogen 

enrichment, biogeochemical 

components of terrestrial and freshwater 

aquatic ecosystems are altered in a way 

that leads to effects on biological 

organisms. 

i. Aquatic Ecosystems 

In freshwater ecosystems, the 

evidence is sufficient to infer a causal 

relationship between nitrogen 

deposition and the alteration of 

biogeochemical cycling in freshwater 

aquatic ecosystems (US EPA, 2008, 

section 3.3.2.3). Nitrogen deposition is 

the main source of nitrogen enrichment 

to headwater streams, lower order 

streams and high elevation lakes. The 

most common chemical indicators that 

were studied included NO3

2¥ and 

dissolved inorganic nitrogen (DIN) 

concentration in surface waters as well 

as the ratio of chlorophyll a to total 

phosphorus. Elevated surface water 

NO3

¥ concentrations occur in both the 

eastern and western U.S. Studies report 

a significant correlation between 

nitrogen deposition and lake 

biogeochemistry by identifying a 

correlation between wet deposition and 

DIN and the ratio of chlorophyll a to 

total phosphate. Recent evidence 

provides examples of lakes and streams 

that are limited by nitrogen and show 

signs of eutrophication in response to 

nitrogen addition. 

The evidence is sufficient to infer a 

causal relationship between nitrogen 

deposition and the alteration of species 

richness, species composition and 

biodiversity in freshwater aquatic 

ecosystems (US EPA, 2008, section 

3.3.5.3). Increased nitrogen deposition 

can cause a shift in community 

composition and reduce algal 

biodiversity, especially in sensitive 

oligotrophic lakes. 

In the ISA, the evidence is sufficient 

to infer a causal relationship between 

nitrogen deposition and the 

biogeochemical cycling of nitrogen and 

carbon in estuaries (US EPA, 2008, 

section 4.3.4.1 and 3.3.2.3). In general, 

estuaries tend to be nitrogen-limited, 

and many currently receive high levels 

of nitrogen input from human activities 

(US EPA, 2009, section 5.1.1). It is 

unknown if atmospheric deposition 

alone is sufficient to cause 

eutrophication; however, the 

contribution of atmospheric nitrogen 

deposition to total nitrogen load is 

calculated for some estuaries and can be 

>40 percent (US EPA, 2009, section 

5.1.1). 

The evidence is sufficient to infer a 

causal relationship between nitrogen 

deposition and the alteration of species 

richness, species composition and 

biodiversity in estuarine ecosystems (US 

EPA, 2008, section 4.3.4.2 and 3.3.5.4). 

Atmospheric and non-atmospheric 

sources of nitrogen contribute to 

increased phytoplankton and algal 

productivity, leading to eutrophication. 

Shifts in community composition, 

reduced hypolimnetic dissolved oxygen 

(DO), decreases in biodiversity, and 

mortality of submerged aquatic 

vegetation are associated with increased 

N deposition in estuarine systems. 

ii. Terrestrial Ecosystems 

The evidence is sufficient to infer a 

causal relationship between nitrogen 

deposition and the alteration of 

biogeochemical cycling in terrestrial 

ecosystems (US EPA, 2008, section 

4.3.1.1 and 3.3.2.1). This is supported 

by numerous observational, deposition 

gradient and field addition experiments 

in sensitive ecosystems. The leaching of 

NO3

¥ in soil drainage waters and the 

export of NO3

¥ in stream water were 

identified as two of the primary 


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indictors of nitrogen enrichment. 

Several nitrogen-addition studies 

indicate that NO3

¥ leaching is induced 

by chronic additions of nitrogen. 

Studies identified in the ISA found that 

surface water NO3

¥ concentrations 

exceeded 1 μeq/L in watersheds 

receiving about 9 to 13 kg N/ha/yr of 

atmospheric nitrogen deposition. 

Nitrogen deposition disrupts the 

nutrient balance of ecosystems with 

numerous biogeochemical effects. The 

chemical indicators that are typically 

measured include NO3

¥ leaching, soil 

C:N ratio, rates of nitrogen 

mineralization, nitrification, 

denitrification, foliar nitrogen 

concentration, and soil water NO3

¥ and 

NH4

∂ concentrations. Note that nitrogen 

saturation (nitrogen leaching from 

ecosystems) does not need to occur to 

cause effects. Substantial leaching of 

NO3

¥ from forest soils to stream water 

can acidify downstream waters, leading 

to effects described in the previous 

section on aquatic acidification. Due to 

the complexity of interactions between 

the nitrogen and carbon cycling, the 

effects of nitrogen on carbon budgets 

(quantified input and output of carbon 

to the ecosystem) are variable. Regional 

trends in net ecosystem productivity 

(NEP) of forests (not managed for 

silviculture) have been estimated 

through models based on gradient 

studies and meta-analysis. Atmospheric 

nitrogen deposition has been shown to 

cause increased litter accumulation and 

carbon storage in above-ground woody 

biomass. In the West, this has lead to 

increased susceptibility to more severe 

fires. Less is known regarding the effects 

of nitrogen deposition on carbon 

budgets of non-forest ecosystems. 

The evidence is sufficient to infer a 

causal relationship between nitrogen 

deposition on the alteration of species 

richness, species composition and 

biodiversity in terrestrial ecosystems 

(US EPA, 2008, section 4.3.1.2). Some 

organisms and ecosystems are more 

sensitive to nitrogen deposition and 

effects of nitrogen deposition are not 

observed in all habitats. The most 

sensitive terrestrial taxa to nitrogen 

deposition are lichens. Empirical 

evidence indicates that lichens in the 

U.S. are affected by deposition levels as 

low as 3 kg N/ha/yr. Alpine ecosystems 

are also sensitive to nitrogen deposition; 

changes in an individual species (Carex 

rupestris) were estimated to occur at 

deposition levels near 4 kg N/ha/yr and 

modeling indicates that deposition 

levels near 10 kg N/ha/yr alter plant 

community assemblages. In several 

grassland ecosystems, reduced species 

diversity and an increase in non-native, 

invasive species are associated with 

nitrogen deposition. 

iii. Ecosystem Sensitivity to Nutrient 

Enrichment 

The numerous ecosystem types that 

occur across the U.S. have a broad range 

of sensitivity to nitrogen deposition (US 

EPA, 2008, Table 4–4). Increased 

deposition to nitrogen-limited 

ecosystems can lead to production 

increases that may be either beneficial 

or adverse depending on the system and 

management goals. 

Organisms in their natural 

environment are commonly adapted to 

a specific regime of nutrient availability. 

Change in the availability of one 

important nutrient, such as nitrogen, 

may result in an imbalance in ecological 

stoichiometry, with effects on ecosystem 

processes, structure and function. In 

general, nitrogen deposition to 

terrestrial ecosystems causes accelerated 

growth rates in some species deemed 

desirable in commercial forests but may 

lead to altered competitive interactions 

among species and nutrient imbalances, 

ultimately affecting biodiversity. The 

onset of these effects occurs with 

nitrogen deposition levels as low as 3 kg 

N/ha/yr in sensitive terrestrial 

ecosystems to nitrogen deposition. In 

aquatic ecosystems, nitrogen that is both 

leached from the soil and directly 

deposited to the water surface can 

pollute the surface water. This causes 

alteration of the diatom community at 

levels as low as 1.5 kg N/ha/yr in 

sensitive freshwater ecosystems. 

The degree of ecosystem effects lies at 

the intersection of nitrogen loading and 

nitrogen-sensitivity. Nitrogen-sensitivity 

is predominately driven by the degree to 

which growth is limited by nitrogen 

availability. Grasslands in the western 

U.S. are typically nitrogen-limited 

ecosystems dominated by a diverse mix 

of perennial forbs and grass species. A 

meta-analysis discussed in the ISA (US 

EPA, 2008, section 3.3.3), indicated that 

nitrogen fertilization increased 

aboveground growth in all non-forest 

ecosystems except for deserts. In other 

words, almost all terrestrial ecosystems 

are nitrogen-limited and will be altered 

by the addition of anthropogenic 

nitrogen. Likewise, a freshwater lake or 

stream must be nitrogen-limited to be 

sensitive to nitrogen-mediated 

eutrophication. There are many 

examples of fresh waters that are 

nitrogen-limited or nitrogen and 

phosphorous (P) co-limited (US EPA, 

2008, section 3.3.3.2). A large dataset 

meta-analysis discussed in the ISA (US 

EPA, 2008, section 3.3.3.2), found that 

nitrogen-limitation occurred as 

frequently as phosphorous-limitation in 

freshwater ecosystems. Additional 

factors that govern the sensitivity of 

ecosystems to nutrient enrichment from 

nitrogen deposition include rates and 

form of nitrogen deposition, elevation, 

climate, species composition, plant 

growth rate, length of growing season, 

and soil nitrogen retention capacity (US 

EPA, 2008, section 4.3). Less is known 

about the extent and distribution of the 

terrestrial ecosystems in the U.S. that 

are most sensitive to the effects of 

nutrient enrichment from atmospheric 

nirogen deposition compared to 

acidification. 

Because the productivity of estuarine 

and near shore marine ecosystems is 

generally limited by the availability of 

nitrogen, they are susceptible to the 

eutrophication effect of nitrogen 

deposition (US EPA, 2008, section 

4.3.4.1). A recent national assessment of 

eutrophic conditions in estuaries found 

the most eutrophic estuaries were 

generally those that had large 

watershed-to-estuarine surface area, 

high human population density, high 

rainfall and runoff, low dilution and 

low flushing rates. In the REA, the 

National Oceanic and Atmospheric 

Administration’s (NOAA) National 

Estuarine Eutrophication Assessment 

(NEEA) assessment tool, Assessment of 

Estuarine Tropic Status (ASSETS) 

categorical Eutrophication Index (EI) 

was used to evaluate eutrophication due 

to atmospheric loading of nitrogen. The 

ASSETS EI is an estimation of the 

likelihood that an estuary is 

experiencing eutrophication or will 

experience eutrophication based on five 

ecological indicators: Chlorophyll a, 

macroalgae, dissolved oxygen, 

nuisance/toxic algal blooms and 

submerged aquatic vegetation (SAV). 

In the REA, two regions were selected 

for case study analysis using ASSETS 

EI, the Chesapeake Bay and Pamlico 

Sound. Both regions received an 

ASSETS EI rating of Bad indicating that 

the estuary had moderate to high 

pressure due to overall human influence 

and a moderate high to high eutrophic 

condition (US EPA, 2009, sections 

5.2.4.1 and 5.2.4.2). These results were 

then considered with SPAtially 

Referenced Regression on Watershed 

Attributes (SPARROW) modeling to 

develop a response curve to examine the 

role of atmospheric nitrogen deposition 

in achieving a desired decrease in load. 

To change the Neuse River Estuary’s EI 

score from Bad to Poor not only must 

100 percent of the total atmospheric 

nitrogen deposition be eliminated, but 

considerably more nitrogen from other 

sources as well must be controlled (US 

EPA, 2009, section 5.2.7.2). In the 

Potomac River estuary, a 78 percent 


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decrease of total nitrogen could move 

the EI score from Bad to Poor (US EPA, 

2009, section 5.2.7.1). The results of this 

analysis indicated decreases in 

atmospheric deposition alone could not 

eliminate coastal eutrophication 

problems due to multiple nonatmospheric nitrogen inputs (US EPA, 

2009, section 7.3.3). However, the 

somewhat arbitrary discreteness of the 

EI scale can mask the benefits of 

decreases in nitrogen between 

categories. 

In general, estuaries tend to be 

nitrogen-limited, and many currently 

receive high levels of nitrogen input 

from human activities to cause 

eutrophication. As reported in the ISA 

(US EPA, 2008, section 3.2.2.2), 

atmospheric nitrogen loads to estuaries 

in the U.S. are estimated to range from 

2 to 8 percent for Guadalupe Bay, Texas 

on the lowest end to as high as 72 

percent for St. Catherines-Sapelo 

estuary, Georgia. The Chesapeake Bay is 

an example of a large, well-studied and 

severely eutrophic estuary that is 

calculated to receive as much as 30 

percent of its total nitrogen load from 

the atmosphere. 

b. Magnitude of Ecosystem Responses 

i. Aquatic Ecosystems 

The magnitude of ecosystem response 

may be thought of on two time scales, 

current conditions and how ecosystems 

have been altered since the onset of 

anthropogenic nitrogen deposition. As 

noted previously, studies found that 

nitrogen-limitation occurs as frequently 

as phosphorous-limitation in freshwater 

ecosystems (US EPA, 2008, section 

3.3.3.2). Recently, a comprehensive 

study of available data from the 

northern hemisphere surveys of lakes 

along gradients of nitrogen deposition 

show increased inorganic nitrogen 

concentration and productivity to be 

correlated with atmospheric nitrogen 

deposition. The results are unequivocal 

evidence of nitrogen limitation in lakes 

with low ambient inputs of nitrogen, 

and increased nitrogen concentrations 

in lakes receiving nitrogen solely from 

atmospheric nitrogen deposition. It has 

been suggested that most lakes in the 

northern hemisphere may have 

originally been nitrogen-limited, and 

that atmospheric nitrogen deposition 

has changed the balance of nitrogen and 

phosphorous in lakes. 

Available data suggest that the 

increases in total nitrogen deposition do 

not have to be large to elicit an 

ecological effect. For example, a 

hindcasting exercise determined that 

the change in Rocky Mountain National 

Park lake algae that occurred between 

1850 and 1964 was associated with an 

increase in wet nitrogen deposition that 

was only about 1.5 kg N/ha. Similar 

changes inferred from lake sediment 

cores of the Beartooth Mountains of 

Wyoming also occurred at about 1.5 kg 

N/ha deposition. Pre-industrial 

inorganic nitrogen deposition is 

estimated to have been only 0.1 to 0.7 

kg N/ha based on measurements from 

remote parts of the world. In the 

western U.S., pre-industrial, or 

background, inorganic nitrogen 

deposition was estimated by to range 

from 0.4 to 0.7 kg N/ha/yr. 

Eutrophication effects from nitrogen 

deposition are most likely to be 

manifested in undisturbed, low nutrient 

surface waters such as those found in 

the higher elevation areas of the western 

U.S. The most severe eutrophication 

from nitrogen deposition effects is 

expected downwind of major urban and 

agricultural centers. High 

concentrations of lake or streamwater 

NO3

¥, indicative of ecosystem 

saturation, have been found at a variety 

of locations throughout the U.S., 

including the San Bernardino and San 

Gabriel Mountains within the Los 

Angeles Air Basin, the Front Range of 

Colorado, the Allegheny mountains of 

West Virginia, the Catskill Mountains of 

New York, the Adirondack Mountains 

of New York, and the Great Smoky 

Mountains in Tennessee (US EPA, 2008, 

section 3.3.8). 

In contrast to terrestrial and 

freshwater systems, atmospheric 

nitrogen load to estuaries contributes to 

the total load but does not necessarily 

drive the effects since other combined 

sources of nitrogen often greatly exceed 

nitrogen deposition. In estuaries, 

nitrogen-loading from multiple 

anthropogenic and non-anthropogenic 

pathways leads to water quality 

deterioration, resulting in numerous 

effects including hypoxic zones, species 

mortality, changes in community 

composition and harmful algal blooms 

that are indicative of eutrophication. 

The following summary is a concise 

overview of the known or anticipated 

effects of nitrogen enrichment on 

estuaries within the U.S. 

There is a scientific consensus (US 

EPA, 2008, section 4.3.4) that nitrogendriven eutrophication in shallow 

estuaries has increased over the past 

several decades and that the 

environmental degradation of coastal 

ecosystems due to nitrogen, 

phosphorus, and other inputs is now a 

widespread occurrence. For example, 

the frequency of phytoplankton blooms 

and the extent and severity of hypoxia 

have increased in the Chesapeake Bay 

and Pamlico estuaries in North Carolina 

and along the continental shelf adjacent 

to the Mississippi and Atchafalaya 

rivers’ discharges to the Gulf of Mexico. 

A recent national assessment of 

eutrophic conditions in estuaries found 

that 65 percent of the assessed systems 

had moderate to high overall eutrophic 

conditions. Most eutrophic estuaries 

occurred in the mid-Atlantic region and 

the estuaries with the lowest degree of 

eutrophication were in the North 

Atlantic. Other regions had mixtures of 

low, moderate, and high degrees of 

eutrophication (US EPA, 2008, section 

4.3.4.3). 

The mid-Atlantic region is the most 

heavily impacted area in terms of 

moderate or high loss of submerged 

aquatic vegetation due to eutrophication 

(US EPA, 2008, section 4.3.4.2). 

Submerged aquatic vegetation is 

important to the quality of estuarine 

ecosystem habitats because it provides 

habitat for a variety of aquatic 

organisms, absorbs excess nutrients, and 

traps sediments (US EPA, 2008, section 

4.3.4.2). It is partly because many 

estuaries and near-coastal marine waters 

are degraded by nutrient enrichment 

that they are highly sensitive to 

potential negative impacts from nitrogen 

addition from atmospheric deposition. 

ii. Terrestrial Ecosystems 

Little is known about the full extent 

and distribution of the terrestrial 

ecosystems in the U.S. that are most 

sensitive to impacts caused by nutrient 

enrichment from atmospheric nitrogen 

deposition. As previously stated, most 

terrestrial ecosystems are nitrogenlimited, therefore they are sensitive to 

perturbation caused by nitrogen 

additions (US EPA, 2008, section 4.3.1). 

Effects are most likely to occur where 

areas of relatively high atmospheric N 

deposition intersect with nitrogenlimited plant communities. The alpine 

ecosystems of the Colorado Front Range, 

chaparral watersheds of the Sierra 

Nevada, lichen and vascular plant 

communities in the San Bernardino 

Mountains and the Pacific Northwest, 

and the southern California coastal sage 

scrub (CSS) community are among the 

most sensitive terrestrial ecosystems. 

There is growing evidence (US EPA, 

2008, section 4.3.1.2) that existing 

grassland ecosystems in the western 

U.S. are being altered by elevated levels 

of N inputs, including inputs from 

atmospheric deposition. 

In the eastern U.S., the degree of 

nitrogen saturation of the terrestrial 

ecosystem is often assessed in terms of 

the degree of NO3

¥ leaching from 

watershed soils into ground water or 

surface water. Studies have estimated 

the number of surface waters at different 


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stages of saturation across several 

regions in the eastern U.S. Of the 85 

northeastern watersheds examined 60 

percent were in Stage 1 or Stage 2 of 

nitrogen saturation on a scale of 0 

(background or pretreatment) to 3 

(visible decline). Of the northeastern 

sites for which adequate data were 

available for assessment, those in Stage 

1 or 2 were most prevalent in the 

Adirondack and Catskill Mountains. 

Effects on individual plant species have 

not been well studied in the U.S. More 

is known about the sensitivity of 

particular plant communities. Based 

largely on results obtained in more 

extensive studies conducted in Europe, 

it is expected that the more sensitive 

terrestrial ecosystems include hardwood 

forests, alpine meadows, arid and semiarid lands, and grassland ecosystems 

(US EPA, 2008, section 3.3.5). 

The REA used published research 

results (US EPA, 2009, section 5.3.1 and 

US EPA, 2008, Table 4.4) to identify 

meaningful ecological benchmarks 

associated with different levels of 

atmospheric nitrogen deposition. These 

are illustrated in Figure 3–4 of the PA. 

The sensitive areas and ecological 

indicators identified by the ISA were 

analyzed further in the REA to create a 

national map that illustrates effects 

observed from ambient and 

experimental atmospheric nitrogen 

deposition loads in relation to 

Community Multi-scale Air Quality 

(CMAQ) 2002 modeling results and 

National Atmospheric Deposition 

Program (NADP) monitoring data. This 

map, reproduced in Figure 3–5 of the 

PA, depicts the sites where empirical 

effects of terrestrial nutrient enrichment 

have been observed and site proximity 

to elevated atmospheric nitrogen 

deposition. 

Based on information in the ISA and 

initial analysis in the REA, further case 

study analyses on terrestrial nutrient 

enrichment of ecosystems were 

developed for the CS community and 

Mixed Conifer Forest (MCF) (US EPA, 

2009). Geographic information systems 

(GIS) analysis supported a qualitative 

review of past field research to identify 

ecological benchmarks associated with 

CSS and mycorrhizal communities, as 

well as MCF nutrient-sensitive 

acidophyte lichen communities, fineroot biomass in Ponderosa pine, and 

leached nitrate in receiving waters. 

The ecological benchmarks that were 

identified for the CSS and the MCF 

communities are included in the suite of 

benchmarks identified in the ISA (US 

EPA, 2008, section 3.3). There are 

sufficient data to confidently relate the 

ecological effect to a loading of 

atmospheric nitrogen. For the CSS 

community, the following ecological 

benchmarks were identified: 

(1) 3.3 kg N/ha/yr—the amount of 

nitrogen uptake by a vigorous stand of 

CSS; above this level, nitrogen may no 

longer be limiting 

(2) 10 kg N/ha/yr—mycorrhizal 

community changes 

For the MCF community, the 

following ecological benchmarks were 

identified: 

(1) 3.1 kg N/ha/yr—shift from sensitive 

to tolerant lichen species 

(2) 5.2 kg N/ha/yr—dominance of the 

tolerant lichen species 

(3) 10.2 kg N/ha/yr—loss of sensitive 

lichen species 

(4) 17 kg N/ha/yr—leaching of nitrate 

into streams. 

These benchmarks, ranging from 3.1 

to 17 kg N/ha/yr, were compared to 

2002 CMAQ/NADP data to discern any 

associations between atmospheric 

deposition and changing communities. 

Evidence supports the finding that 

nitrogen alters CSS and MCF 

communities. Key findings include the 

following: 2002 CMAQ/NADP nitrogen 

deposition data show that the 3.3 kg N/ 

ha/yr benchmark has been exceeded in 

more than 93 percent of CSS areas 

(654,048 ha). These deposition levels 

are a driving force in the degradation of 

CSS communities. Although CSS 

decline has been observed in the 

absence of fire, the contributions of 

deposition and fire to the CSS decline 

require further research. The CSS is 

fragmented into many small parcels, 

and the 2002 CMAQ/NADP 12-km grid 

data are not fine enough to fully validate 

the relationship between CSS 

distribution, nitrogen deposition, and 

fire. The 2002 CMAQ/NADP nitrogen 

deposition data exceeds the 3.1 kg N/ha/ 

yr benchmark in more than 38 percent 

(1,099,133 ha) of MCF areas, and nitrate 

leaching has been observed in surface 

waters. Ozone effects confound nitrogen 

effects on MCF acidophyte lichen, and 

the interrelationship between fire and 

nitrogen cycling requires additional 

research. 

c. Key Uncertainties Associated With 

Nutrient Enrichment 

There are different levels of 

uncertainty associated with 

relationships between deposition, 

ecological effects and ecological 

indicators. The criteria used in the REA 

to evaluate the degree of confidence in 

the data, modeling and ecological effect 

function are detailed in chapter 7 of the 

REA. Below is a discussion of 

uncertainty relating aquatic and 

terrestrial ecosystems to nutrient 

enrichment effects. 

i. Aquatic Ecosystems 

The approach for assessing 

atmospheric contributions to total 

nitrogen loading in the REA was to 

consider the main-stem river to an 

estuary (including the estuary) rather 

than an entire estuary system or bay. 

The biological indicators used in the 

NOAA ASSETS EI required the 

evaluation of many national databases 

including the US Geological Survey 

National Water Quality Assessment 

(NAWQA) files, EPA’s STORage and 

RETrieval (STORET) database, NOAA’s 

Estuarine Drainage Areas data and 

EPA’s water quality standards nutrient 

criteria for rivers and lakes (US EPA, 

2009, Appendix 6 and Table 1.2.–1). 

Both the SPARROW modeling for 

nitrogen loads and assessment of 

estuary conditions under NOAA 

ASSETS EI, have been applied on a 

national scale. The REA concludes that 

the available data are medium quality 

with intermediate confidence about the 

use of these data and their values for 

extrapolating to a larger regional area 

(US EPA, 2009, section 7.3.1). 

Intermediate confidence is associated 

with the modeling approach using 

ASSETS EI and SPARROW. The REA 

states there is low confidence with the 

ecological effect function due to the 

results of the analysis which indicated 

that reductions in atmospheric 

deposition alone could not solve coastal 

eutrophication problems due to 

multiple non-atmospheric nitrogen 

inputs (US EPA, 2009, section 7.3.3). 

ii. Terrestrial Ecosystems 

Ecological thresholds are identified 

for CSS and MCF areas and these data 

are considered to be of high quality, 

however, the ability to extrapolate these 

data to larger regional areas is limited 

(US EPA, 2009, section 7.4.1). No 

quantitative modeling was conducted or 

ecological effect function developed for 

terrestrial nutrient enrichment reflecting 

the uncertainties associated with these 

depositional effects. 

4. Other Ecological Effects 

It is stated in the ISA (US EPA, 2008, 

section 3.4.1 and 4.5) that mercury is a 

highly neurotoxic contaminant that 

enters the food web as a methylated 

compound, methylmercury (MeHg). 

Mercury is principally methylated by 

sulfur-reducing bacteria and can be 

taken up by microorganisms, 

zooplankton and macroinvertebrates. 

The contaminant is concentrated in 

higher trophic levels, including fish 

eaten by humans. Experimental 

evidence has established that only 

inconsequential amounts of MeHg can 


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be produced in the absence of sulfate. 

Once MdHg is present, other variables 

influence how much accumulates in 

fish, but elevated mercury levels in fish 

can only occur where substantial 

amounts of MeHg are present. Current 

evidence indicates that in watersheds 

where mercury is present, increased 

oxides of sulfur deposition very likely 

results in additional production of 

MeHg which leads to greater 

accumulation of MeHg concentrations 

in fish. With respect to sulfur deposition 

and mercury methylation, the final ISA 

determined that ‘‘[t]he evidence is 

sufficient to infer a causal relationship 

between sulfur deposition and increased 

mercury methylation in wetlands and 

aquatic environments.’’ 

The production of meaningful 

amounts of MeHg requires the presence 

of SO4

2¥ and mercury, and where 

mercury is present, increased 

availability of SO4

2¥ results in 

increased production of MeHg. There is 

increasing evidence on the relationship 

between sulfur deposition and increased 

methylation of mercury in aquatic 

environments; this effect occurs only 

where other factors are present at levels 

within a range to allow methylation. 

The production of MeHg requires the 

presence of SO4

2¥ and mercury, but the 

amount of MeHg produced varies with 

oxygen content, temperature, pH, and 

supply of labile organic carbon (US 

EPA, 2008, section 3.4). In watersheds 

where changes in sulfate deposition did 

not produce an effect, one or several of 

those interacting factors were not in the 

range required for meaningful 

methylation to occur (US EPA, 2008, 

section 3.4). Watersheds with 

conditions known to be conducive to 

mercury methylation can be found in 

the northeastern U.S. and southeastern 

Canada. 

While the relationship between sulfur 

and MeHg production was concluded to 

be causal in the ISA, the REA concluded 

that there was insufficient evidence to 

quantify the relationship between sulfur 

and MeHg. Therefore only a qualitative 

assessment was included in chapter 6 of 

the REA. The PA was then unable to 

make a determination as to the 

adequacy of the existing SO2 standards 

in protecting against welfare effects 

associated with increased mercury 

methylation. 

B. Risk and Exposure Assessment 

The risk and exposure assessment 

conducted for the current review was 

developed to describe potential risk 

from current and future deposition of 

oxides of nitrogen and sulfur to 

sensitive ecosystems. The case study 

analyses in the REA show that there is 

confidence that known or anticipated 

adverse ecological effects are occurring 

under current ambient loadings of 

nitrogen and sulfur in sensitive 

ecosystems across the U.S. An overview 

of the material covered in the REA, a 

summary of the key findings from the 

air quality analyses, acidification and 

nutrient enrichment case studies, and 

general conclusions from evaluating 

additional welfare effects, are presented 

below. 

1. Overview of the Risk and Exposure 

Assessment 

The REA evaluates the relationships 

between atmospheric concentrations, 

deposition, biologically relevant 

exposures, targeted ecosystem effects, 

and ecosystem services. To evaluate the 

nature and magnitude of adverse effects 

associated with deposition, the REA 

also examines various ways to quantify 

the relationships between air quality 

indicators, deposition of biologically 

available forms of nitrogen and sulfur, 

ecologically relevant indicators relating 

to deposition, exposure and effects on 

sensitive receptors, and related effects 

resulting in changes in ecosystem 

structure and services. The intent is to 

determine the exposure metrics that 

incorporate the temporal considerations 

(i.e., biologically relevant timescales), 

pathways, and ecologically relevant 

indicators necessary to determine the 

effects on these ecosystems. To the 

extent feasible, the REA evaluates the 

overall load to the system for nitrogen 

and sulfur, as well as the variability in 

ecosystem responses to these pollutants. 

It also evaluates the contributions of 

atmospherically deposited nitrogen and 

sulfur individually relative to the 

combined atmospheric loadings of both 

elements together.. Since oxidized 

nitrogen is the listed criteria pollutant 

(currently measured by the ambient air 

quality indicator NO2) for the 

atmospheric contribution to total 

nitrogen, the REA examines the 

contribution of nitrogen oxides to total 

reactive nitrogen in the atmosphere, 

relative to the contributions of reduced 

forms of nitrogen (e.g., ammonia, 

ammonium), to ultimately assess how a 

meaningful secondary NAAQS might be 

structured. 

The REA focuses on ecosystem 

welfare effects that result from the 

deposition of total reactive nitrogen and 

sulfur. Because ecosystems are diverse 

in biota, climate, geochemistry, and 

hydrology, response to pollutant 

exposures can vary greatly between 

ecosystems. In addition, these diverse 

ecosystems are not distributed evenly 

across the United States. To target 

nitrogen and sulfur acidification and 

nitrogen and sulfur enrichment, the 

REA addresses four main targeted 

ecosystem effects on terrestrial and 

aquatic systems identified by the ISA 

(US EPA, 2008): Aquatic acidification 

due to nitrogen and sulfur; terrestrial 

acidification due to nitrogen and sulfur; 

aquatic nutrient enrichment, including 

eutrophication; and terrestrial nutrient 

enrichment. 

In addition to these four targeted 

ecosystem effects, the REA also 

qualitatively addresses the influence of 

sulfur oxides deposition on MeHg 

production; nitrous oxide (N2O) effects 

on climate; nitrogen effects on primary 

productivity and biogenic greenhouse 

gas (GHG) fluxes; and phytotoxic effects 

on plants. 

Because the targeted ecosystem effects 

outlined above are not evenly 

distributed across the U.S., the REA 

identified case studies for each targeted 

effects based on ecosystems identified 

as sensitive to nitrogen and/or sulfur 

deposition effects. Eight case study 

areas and two supplemental study areas 

(Rocky Mountain National Park and 

Little Rock Lake, Wisconsin) are 

summarized in the REA based on 

ecosystem characteristics, indicators, 

and ecosystem service information. Case 

studies selected for aquatic acidification 

effects were the Adirondack Mountains 

and Shenandoah National Park. Kane 

Experimental Forest in Pennsylvania 

and Hubbard Brook Experimental Forest 

in New Hampshire were selected as case 

studies for terrestrial acidification. 

Aquatic nutrient enrichment case study 

locations were selected in the Potomac 

River Basin upstream of Chesapeake Bay 

and the Neuse River Basin upstream of 

the Pamlico Sound in North Carolina. 

The CSS communities in southern 

California and the MCF communities in 

the San Bernardino and Sierra Nevada 

Mountains of California were selected as 

case studies for terrestrial nutrient 

enrichment. Two supplemental areas 

were also chosen, one in Rocky 

Mountain National Park for terrestrial 

nutrient enrichment and one in Little 

Rock Lake, Wisconsin for aquatic 

nutrient enrichment. 

2. Key Findings 

In summary, based on case study 

analyses, the REA concludes that known 

or anticipated adverse ecological effects 

are occurring under current conditions 

and further concludes that these adverse 

effects continue into the future. Key 

findings from the air quality analyses, 

acidification and nutrient enrichment 

case studies, as well as general 

conclusions from evaluating additional 

welfare effects, are summarized below. 


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a. Air Quality Analyses 

The air quality analyses in the REA 

encompass the current emissions 

sources of nitrogen and sulfur, as well 

as atmospheric concentrations, 

estimates of deposition of total nitrogen, 

policy-relevant background, and nonatmospheric loadings of nitrogen and 

sulfur to ecosystems, both nationwide 

and in the case study areas. Spatial 

fields of deposition were created using 

wet deposition measurements from the 

NADP National Trends Network and dry 

deposition predictions from the 2002 

CMAQ model simulation. Some key 

conclusions from this analysis are: 

(1) Total reactive nitrogen deposition 

and sulfur deposition are much greater 

in the East compared to most areas of 

the West. 

(2) These regional differences in 

deposition correspond to the regional 

differences in oxides of nitrogen and 

SO2 concentrations and emissions, 

which are also higher in the East. 

Oxides of nitrogen emissions are much 

greater and generally more widespread 

than NH3 emissions nationwide; high 

NH3 emissions tend to be more local 

(e.g., eastern North Carolina) or subregional (e.g., the upper Midwest and 

Plains states). The relative amounts of 

oxidized versus reduced nitrogen 

deposition are consistent with the 

relative amounts of oxides of nitrogen 

and NH3 emissions. Oxidized nitrogen 

deposition exceeds reduced nitrogen 

deposition in most of the case study 

areas; the major exception being the 

Neuse River/Neuse River Estuary Case 

Study Area. 

(3) Reduced nitrogen deposition 

exceeds oxidized nitrogen deposition in 

the vicinity of local sources of NH3. 

(4) There can be relatively large 

spatial variations in both total reactive 

nitrogen deposition and sulfur 

deposition within a case study area; this 

occurs particularly in those areas that 

contain or are near a high emissions 

source of oxides of nitrogen, NH3 and/ 

or SO2. 

(5) The seasonal patterns in 

deposition differ between the case study 

areas. For the case study areas in the 

East, the season with the greatest 

amounts of total reactive nitrogen 

deposition correspond to the season 

with the greatest amounts of sulfur 

deposition. Deposition peaks in spring 

in the Adirondack, Hubbard Brook 

Experimental Forest, and Kane 

Experimental Forest case study areas, 

and it peaks in summer in the Potomac 

River/Potomac Estuary, Shenandoah, 

and Neuse River/Neuse River Estuary 

case study areas. For the case study 

areas in the West, there is less 

consistency in the seasons with greatest 

total reactive nitrogen and sulfur 

deposition in a given area. In general, 

both nitrogen and/or sulfur deposition 

peaks in spring or summer. The 

exception to this is the Sierra Nevada 

Range portion of the MCF Case Study 

Area, in which sulfur deposition is 

greatest in winter. 

b. Deposition-Related Aquatic 

Acidification 

The role of aquatic acidification in 

two eastern United States areas— 

northeastern New York’s Adirondack 

area and the Shenandoah area in 

Virginia—was analyzed in the REA to 

assess surface water trends in SO4

2¥

and NO3

¥ concentrations and ANC 

levels and to affirm the understanding 

that reductions in deposition could 

influence the risk of acidification. 

Monitoring data from the EPAadministered TIME)/LTM programs and 

the EMAP were assessed for the years 

1990 to 2006, and past, present and 

future water quality levels were 

estimated using both steady-state and 

dynamic biogeochemical models. 

Although wet deposition rates for SO2

and oxides of nitrogen in the 

Adirondack Case Study Area have 

reduced since the mid-1990s, current 

concentrations are still well above preacidification (1860) conditions. The 

MAGIC modeling predicts NO3

¥ and 

SO4

2¥ are 17- and 5-fold higher today, 

respectively. The estimated average 

ANC for 44 lakes in the Adirondack 

Case Study Area is 62.1 μeq/L (±15.7 

μeq/L); 78 percent of all monitored lakes 

in the Adirondack Case Study Area have 

a current risk of Elevated, Severe, or 

Acute. Of the 78 percent, 31 percent 

experience episodic acidification, and 

18 percent are chronically acidic today. 

(1) Based on the steady-state critical 

load model for the year 2002, 18 

percent, 28 percent, 44 percent, and 58 

percent of 169 modeled lakes received 

combined total sulfur and nitrogen 

deposition that exceeded critical loads 

corresponding to ANC limits of 0, 20, 

50, and 100 μeq/L respectively. 

(2) Based on a deposition scenario 

that maintains current emission levels 

to 2020 and 2050, the simulation 

forecast indicates no improvement in 

water quality in the Adirondack Case 

Study Area. The percentage of lakes 

within the Elevated to Acute Concern 

classes remains the same in 2020 and 

2050. 

(3) Since the mid-1990s, streams in 

the Shenandoah Case Study Area have 

shown slight declines in NO3 and SO4

2¥

concentrations in surface waters. The 

ANC levels increased from about 50 

μeq/L in the early 1990s to >75 μeq/L 

until 2002, when ANC levels declined 

back to 1991–1992 levels. Current 

concentrations are still above preacidification (1860) conditions. The 

MAGIC modeling predicts surface water 

concentrations of NO3 and SO4

2¥ are 

10- and 32-fold higher today, 

respectively. The estimated average 

ANC for 60 streams in the Shenandoah 

Case Study Area is 57.9 μeq/L (±4.5 μeq/ 

L). Fifty-five percent of all monitored 

streams in the Shenandoah Case Study 

Area have a current risk of Elevated, 

Severe, or Acute. Of the 55 percent, 18 

percent experience episodic 

acidification, and 18 percent are 

chronically acidic today. 

(4) Based on the steady-state critical 

load model for the year 2002, 52 

percent, 72 percent, 85 percent and 93 

percent of 60 modeled streams received 

combined total sulfur and nitrogen 

deposition that exceeded critical loads 

corresponding to ANC limits of 0, 20, 

50, and 100 μeq/L respectively. 

(5) Based on a deposition scenario 

that maintains current emission levels 

to 2020 and 2050, the simulation 

forecast indicates that a large number of 

streams would still have Elevated to 

Acute problems with acidity. 

c. Deposition-Related Terrestrial 

Acidification 

The role of terrestrial acidification 

was examined in the REA using a 

critical load analysis for sugar maple 

and red spruce forests in the eastern 

U.S. by using the BC/Al ratio in 

acidified forest soils as an indicator to 

assess the impact of nitrogen and sulfur 

deposition on tree health. These are the 

two most commonly studied species in 

North America for impacts of 

acidification. At a BC/Al ratio of 1.2, red 

spruce growth can be reduced by 20 

percent. Sugar maple growth can be 

reduced by 20 percent at a BC/Al ratio 

of 0.6. Key findings of the case study are 

summarized below. 

(1) Case study results suggest that the 

health of at least a portion of the sugar 

maple and red spruce growing in the 

U.S. may have been compromised with 

acidifying total nitrogen and sulfur 

deposition in 2002. The 2002 CMAQ/ 

NADP total nitrogen and sulfur 

deposition levels exceeded three 

selected critical loads in 3 percent to 75 

percent of all sugar maple plots across 

24 states. The three critical loads ranged 

from 6,008 to 107 eq/ha/yr for the BC/ 

Al ratios of 0.6, 1.2, and 10.0 (increasing 

levels of tree protection). The 2002 

CMAQ/NADP total nitrogen and sulfur 

deposition levels exceeded three 

selected critical loads in 3 percent to 36 

percent of all red spruce plots across 

eight states. The three critical loads 


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ranged from 4,278 to 180 eq/ha/yr for 

the Bc/Al ratios of 0.6, 1.2, and 10.0 

(increasing levels of tree protection). 

(2) The SMB model assumptions 

made for base cation weathering (Bcw) 

and forest soil ANC input parameters 

are the main sources of uncertainty 

since these parameters are rarely 

measured and require researchers to use 

default values. 

(3) The pattern of case study results 

suggests that nitrogen and sulfur 

acidifying deposition in the sugar maple 

and red spruce forest areas studied were 

similar in magnitude to the critical 

loads for those areas and both 

ecosystems are likely to be sensitive to 

any future changes in the levels of 

deposition. 

d. Deposition-Related Aquatic Nutrient 

Enrichment 

The role of nitrogen deposition in two 

main stem rivers feeding their 

respective estuaries was analyzed in the 

REA to determine if decreases in 

deposition could influence the risk of 

eutrophication as predicted using the 

ASSETS EI scoring system in tandem 

with SPARROW modeling. This 

modeling approach provides a 

transferrable, intermediate-level 

analysis of the linkages between 

atmospheric deposition and receiving 

waters, while providing results on 

which conclusions could be drawn. A 

summary of findings follows: 

(1) The 2002 CMAQ/NADP results 

showed that an estimated 40,770,000 

kilograms (kg) of total nitrogen was 

deposited in the Potomac River 

watershed. The SPARROW modeling 

predicted that 7,380,000 kg N/yr of the 

deposited nitrogen reached the estuary 

(20 percent of the total load to the 

estuary). The overall ASSETS EI for the 

Potomac River and Potomac Estuary was 

Bad (based on all sources of N). 

(2) To improve the Potomac River and 

Potomac Estuary ASSETS EI score from 

Bad to Poor, a decrease of at least 78 

percent in the 2002 total nitrogen 

atmospheric deposition load to the 

watershed would be required. 

(3) The 2002 CMAQ/NADP results 

showed that an estimated 18,340,000 kg 

of total nitrogen was deposited in the 

Neuse River watershed. The SPARROW 

modeling predicted that 1,150,000 kg N/ 

yr of the deposited nitrogen reached the 

estuary (26 percent of the total load to 

the estuary). The overall ASSETS EI for 

the Neuse River/Neuse River Estuary 

was Bad. 

(4) It was found that the Neuse River/ 

Neuse River Estuary ASSETS EI score 

could not be improved from Bad to Poor 

with decreases only in the 2002 

atmospheric deposition load to the 

watershed. Additional reductions would 

be required from other nitrogen sources 

within the watershed. 

The small effect of decreasing 

atmospheric deposition in the Neuse 

River watershed is because the other 

nitrogen sources within the watershed 

are more influential than atmospheric 

deposition in affecting the total nitrogen 

loadings to the Neuse River Estuary, as 

estimated with the SPARROW model. A 

water body’s response to nutrient 

loading depends on the magnitude (e.g., 

agricultural sources have a higher 

influence in the Neuse than in the 

Potomac), spatial distribution, and other 

characteristics of the sources within the 

watershed; therefore a reduction in 

nitrogen deposition does not always 

produce a linear response in reduced 

load to the estuary, as demonstrated by 

these two case studies. 

e. Deposition-Related Terrestrial 

Nutrient Enrichment 

California CSS and MCF communities 

were the focus of the Terrestrial 

Nutrient Enrichment Case Studies of the 

REA. Geographic information systems 

analysis supported a qualitative review 

of past field research to identify 

ecological benchmarks associated with 

CSS and mycorrhizal communities, as 

well as MCF’s nutrient-sensitive 

acidophyte lichen communities, fineroot biomass in Ponderosa pine and 

leached nitrate in receiving waters. 

These benchmarks, ranging from 3.1 to 

17 kg N/ha/yr, were compared to 2002 

CMAQ/NADP data to discern any 

associations between atmospheric 

deposition and changing communities. 

Evidence supports the finding that 

nitrogen alters CSS and MCF. Key 

findings include the following: 

(1) The 2002 CMAQ/NADP nitrogen 

deposition data show that the 3.3 kg N/ 

ha/yr benchmark has been exceeded in 

more than 93 percent of CSS areas 

(654,048 ha). This suggests that such 

deposition is a driving force in the 

degradation of CSS communities. One 

potentially confounding factor is the 

role of fire. Although CSS decline has 

been observed in the absence of fire, the 

contributions of deposition and fire to 

the CSS decline require further research. 

The CSS is fragmented into many small 

parcels, and the 2002 CMAQ/NADP 12- 

km grid data are not fine enough to fully 

validate the relationship between CSS 

distribution, nitrogen deposition, and 

fire. 

(2) The 2002 CMAQ/NADP nitrogen 

deposition data exceeds the 3.1 kg N/ha/ 

yr benchmark in more than 38% 

(1,099,133 ha) of MCF areas, and nitrate 

leaching has been observed in surface 

waters. Ozone effects confound nitrogen 

effects on MCF acidophyte lichen, and 

the interrelationship between fire and 

nitrogen cycling requires additional 

research. 

f. Additional Effects 

Ecological effects have also been 

documented across the U.S. where 

elevated nitrogen deposition has been 

observed, including the eastern slope of 

the Rocky Mountains where shifts in 

dominant algal species in alpine lakes 

have occurred where wet nitrogen 

deposition was only about 1.5 kg N/ha/ 

yr. High alpine terrestrial communities 

have a low capacity to sequester 

nitrogen deposition, and monitored 

deposition exceeding 3 to 4 kg N/ha/yr 

could lead to community-level changes 

in plant species, lichens and 

mycorrhizae. 

Additional welfare effects are 

documented, but examined less 

extensively, in the REA. These effects 

include qualitative discussions related 

to visibility and materials damage, such 

as corrosion, erosion, and soiling of 

paint and buildings which are being 

addressed in the PM NAAQS review 

currently underway. A discussion of the 

causal relationship between sulfur 

deposition (as sulfate, SO4

2¥) and 

increased mercury methylation in 

wetlands and aquatic environments is 

also included in the REA. On this 

subject the REA concludes that 

decreases in SO4

2¥ deposition will 

likely result in decreases in MeHg 

concentration; however, spatial and 

biogeochemical variations nationally 

hinder establishing large scale doseresponse relationships. 

Several additional issues concerning 

oxides of nitrogen were addressed in the 

REA. Consideration was also given to 

N2O, a potent GHG. The REA concluded 

that it is most appropriate to analyze the 

role of N2O in the context of all of the 

GHGs rather than as part of the REA for 

this review. The REA considered 

nitrogen deposition and its correlation 

with the rate of photosynthesis and net 

primary productivity. Nitrogen addition 

ranging from 15.4 to 300 kg N/ha/yr is 

documented as increasing wetland N2O 

production by an average of 207 percent 

across all ecosystems. Nitrogen addition 

ranging from 30 to 240 kg N/ha/yr 

increased CH4 emissions by 115 percent, 

averaged across all ecosystems, and 

methane uptake was reduced by 38 

percent averaged across all ecosystems 

when nitrogen addition ranged from 10 

to 560 kg N/ha/yr, but reductions were 

only significant for coniferous and 

deciduous forests. The heterogeneity of 

ecosystems across the U.S., however, 

introduces variations into dose-response 

relationships. 


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The phytotoxic effects of oxides of 

nitrogen and sulfur on vegetation were 

also briefly discussed in the REA which 

concluded that since a unique 

secondary NAAQS exists for SO2, and 

concentrations of nitric oxide (NO), NO2

and PAN are rarely high enough to have 

phytotoxic effects on vegetation, further 

assessment was not warranted at this 

time. 

3. Conclusions on Effects 

For aquatic and terrestrial 

acidification effects, a similar 

conceptual approach was used (critical 

loads) to evaluate the impacts of 

multiple pollutants on an ecological 

endpoint, whereas the approaches used 

for aquatic and terrestrial nutrient 

enrichment were fundamentally 

distinct. Although the ecological 

indicators for aquatic and terrestrial 

acidification (i.e., ANC and BC/Al) are 

very different, both ecological indicators 

are well-correlated with effects such as 

reduced biodiversity and growth. While 

aquatic acidification is clearly the 

targeted effect area with the highest 

level of confidence, the relationship 

between atmospheric deposition and an 

ecological indicator is also quite strong 

for terrestrial acidification. The main 

drawback with the understanding of 

terrestrial acidification is that the data 

are based on laboratory responses rather 

than field measurements. Other 

stressors that are present in the field but 

that are not present in the laboratory 

may confound this relationship. 

For nutrient enrichment effects, the 

REA utilized different types of 

indicators for aquatic and terrestrial 

effects to assess both the likelihood of 

adverse effects to ecosystems and the 

relationship between adverse effects and 

atmospheric sources of oxides of 

nitrogen. The ecological indicator 

chosen for aquatic nutrient enrichment, 

the ASSETS EI, seems to be inadequate 

to relate atmospheric deposition to the 

targeted ecological effect, likely due to 

the many other confounding factors. 

Further, there is far less confidence 

associated with the understanding of 

aquatic nutrient enrichment because of 

the large contributions from nonatmospheric sources of nitrogen and the 

influence of both oxidized and reduced 

forms of nitrogen, particularly in large 

watersheds and coastal areas. However, 

a strong relationship exists between 

atmospheric deposition of nitrogen and 

ecological effects in high alpine lakes in 

the Rocky Mountains because 

atmospheric deposition is the only 

source of nitrogen to these systems. 

There is also a strong weight-ofevidence regarding the relationships 

between ecological effects attributable to 

terrestrial nitrogen nutrient enrichment; 

however, ozone and climate change may 

be confounding factors. In addition, the 

response for other species or species in 

other regions of the U.S. has not been 

quantified. 

C. Adversity of Effects to Public Welfare 

Characterizing a known or anticipated 

adverse effect to public welfare is an 

important component of developing any 

secondary NAAQS. According to the 

CAA, welfare effects include: ‘‘Effects 

on soils, water, crops, vegetation, 

manmade materials, animals, wildlife, 

weather, visibility, and climate, damage 

to and deterioration of property, and 

hazards to transportation, as well as 

effect on economic values and on 

personal comfort and well-being, 

whether caused by transformation, 

conversion, or combination with other 

air pollutants’’ (CAA, Section 302(h)). 

While the text above lists a number of 

welfare effects, these effects do not 

define public welfare in and of 

themselves. 

Although there is no specific 

definition of adversity to public welfare, 

the paradigm of linking adversity to 

public welfare to disruptions in 

ecosystem structure and function has 

been used broadly by EPA to categorize 

effects of pollutants from the cellular to 

the ecosystem level. An evaluation of 

adversity to public welfare might 

consider the likelihood, type, 

magnitude, and spatial scale of the 

effect as well as the potential for 

recovery and any uncertainties relating 

to these considerations. 

Similar concepts were used in past 

reviews of secondary NAAQS for ozone 

and PM (relating to visibility), as well as 

in initial reviews of effects from lead 

deposition. Because oxides of nitrogen 

and sulfur are deposited from ambient 

sources into ecosystems where they 

affect changes to organisms, populations 

and ecosystems, the concept of 

adversity to public welfare as a result of 

alterations in structure and function of 

ecosystems is an appropriate 

consideration for this review. 

Based on information provided in the 

PA, the following section discusses how 

ecological effects from deposition of 

oxides of nitrogen and sulfur relate to 

adversity to public welfare. In the PA, 

public welfare was discussed in terms of 

loss of ecosystem services (defined 

below), which in some cases can be 

monetized. Each of the four main effect 

areas (aquatic and terrestrial 

acidification and aquatic and terrestrial 

nutrient over-enrichment) are discussed 

including current ecological effects and 

associated ecosystem services. 

1. Ecosystem Services 

The PA defines ecosystem services as 

the benefits individuals and 

organizations obtain from ecosystems. 

Ecosystem services can be classified as 

provisioning (food and water), 

regulating (control of climate and 

disease), cultural (recreational, 

existence, spiritual, educational), and 

supporting (nutrient cycling). 

Conceptually, changes in ecosystem 

services may be used to aid in 

characterizing a known or anticipated 

adverse effect to public welfare. In the 

REA and PA ecosystem services are 

discussed as a method of assessing the 

magnitude and significance to the 

public of resources affected by ambient 

concentrations of oxides of nitrogen and 

sulfur and deposition in sensitive 

ecosystems. 

The EPA has in previous NAAQS 

reviews defined ecological goods and 

services for the purposes of a Regulatory 

Impact Analysis as the ‘‘outputs of 

ecological functions or processes that 

directly or indirectly contribute to social 

welfare or have the potential to do so in 

the future. Some outputs may be bought 

and sold, but most are not marketed.’’ It 

is especially important to acknowledge 

that it is difficult to measure and/or 

monetize the goods and services 

supplied by ecosystems. It can be 

informative in characterizing adversity 

to public welfare to attempt to place an 

economic valuation on the set of goods 

and services that have been identified 

with respect to a change in policy; 

however it must be noted that this 

valuation will be incomplete and 

illustrative only. 

Knowledge about the relationships 

linking ambient concentrations and 

ecosystem services is considered in the 

PA as one method by which to inform 

a policy judgment on a known or 

anticipated adverse public welfare 

effect. For example, a change in an 

ecosystem structure and process, such 

as foliar injury, would be classified as 

an ecological effect, with the associated 

changes in ecosystem services, such as 

primary productivity, food availability, 

forest products, and aesthetics (e.g., 

scenic viewing), classified as public 

welfare effects. Additionally, changes in 

biodiversity would be classified as an 

ecological effect, and the associated 

changes in ecosystem services— 

productivity, existence (nonuse) value, 

recreational viewing and aesthetics— 

would also be classified as public 

welfare effects. 

As described in chapters 4 and 5 of 

the REA, case study analyses were 

performed that link deposition in 

sensitive ecosystems to changes in a 


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given ecological indicator (e.g., for 

aquatic acidification, to changes in 

ANC) and then to changes in 

ecosystems. Appendix 8 of the REA 

links the changes in ecosystems to the 

services they provide (e.g., fish species 

richness and its influence on 

recreational fishing). To the extent 

possible for each targeted effect area, the 

REA linked ambient concentrations of 

nitrogen and sulfur (i.e., ambient air 

quality indicators) to deposition in 

sensitive ecosystems (i.e., exposure 

pathways), and then to system response 

as measured by a given ecological 

indicator (e.g., lake and stream 

acidification as measured by ANC). The 

ecological effect (e.g., changes in fish 

species richness) was then, where 

possible, associated with changes in 

ecosystem services and the 

corresponding public welfare effects 

(e.g., recreational fishing). 

2. Effects on Ecosystem Services 

The process used to link ecological 

indicators to ecosystem services is 

discussed extensively in appendix 8 of 

the REA. In brief, for each case study 

area assessed, the ecological indicators 

are linked to an ecological response that 

is subsequently linked to associated 

services to the extent possible. For 

example, in the case study for aquatic 

acidification the chosen ecological 

indicator is ANC which can be linked to 

the ecosystem service of recreational 

fishing. Although recreational fishing 

losses are the only service effects that 

can be independently quantified or 

monetized at this time, there are 

numerous other ecosystem services that 

may be related to the ecological effects 

of acidification. 

While aquatic acidification is the 

focus of this proposed standard, the 

other effect areas were also analyzed in 

the REA and these ecosystems are being 

harmed by nitrogen and sulfur 

deposition and will obtain some 

measure of protection with any decrease 

in that deposition regardless of the 

reason for the decrease. The following 

summarizes the current levels of 

specific ecosystem services for aquatic 

and terrestrial acidification and aquatic 

and terrestrial nutrient over-enrichment 

and attempts to quantify and when 

possible monetize the harm to public 

welfare, as represented by ecosystem 

services, due to nitrogen and sulfur 

deposition. 

a. Aquatic Acidification 

Acidification of aquatic ecosystems 

primarily affects the ecosystem services 

that are derived from the fish and other 

aquatic life found in surface waters. In 

the northeastern United States, the 

surface waters affected by acidification 

are not a major source of commercially 

raised or caught fish; however, they are 

a source of food for some recreational 

and subsistence fishers and for other 

consumers. Although data and models 

are available for examining the effects 

on recreational fishing, relatively little 

data are available for measuring the 

effects on subsistence and other 

consumers. Inland waters also provide 

aesthetic and educational services along 

with non-use services, such as existence 

value (protection and preservation with 

no expectation of direct use). In general, 

inland surface waters such as lakes, 

rivers, and streams also provide a 

number of regulating services, playing a 

role in hydrological regimes and climate 

regulation. There is little evidence that 

acidification of freshwaters in the 

northeastern U.S. has significantly 

degraded these specific services; 

however, freshwater ecosystems also 

provide biological control services by 

providing environments that sustain 

delicate aquatic food chains. The toxic 

effects of acidification on fish and other 

aquatic life impair these services by 

disrupting the trophic structure of 

surface waters. Although it is difficult to 

quantify these services and how they are 

affected by acidification, it is worth 

noting that some of these services may 

be captured through measures of 

provisioning and cultural services. For 

example, these biological control 

services may serve as ‘‘intermediate’’ 

inputs that support the production of 

‘‘final’’ recreational fishing and other 

cultural services. 

As summarized in Chapter 4 of the 

PA, recent studies indicate that 

acidification of lakes and streams can 

result in significant loss in economic 

value. For example, data indicate that 

more than 9 percent of adults in the 

northeastern part of the country 

participate annually in freshwater 

fishing yielding 140 million freshwater 

fishing days. Each fishing day has an 

estimated average value per day of $35. 

Therefore, the implied total annual 

value of freshwater fishing in the 

northeastern U.S. was $5 billion in 

2006. Embedded in these numbers is a 

degree of harm to recreational fishing 

services due to acidification that has 

occurred over time. These harms have 

not been quantified on a regional scale; 

however, a case study was conducted in 

the Adirondacks area (US EPA, 2011, 

section 4.4.2). 

In the Adirondacks case study, 

estimates of changes in recreational 

fishing services were determined, as 

well as changes more broadly in 

‘‘cultural’’ ecosystem services 

(including recreational, aesthetic, and 

nonuse services). First, the MAGIC 

model (US EPA, 2009, Appendix 8 and 

section 2.2) was applied to 44 lakes to 

predict what ANC levels would be 

under both ‘‘business as usual’’ 

conditions (i.e., allowing for some 

decline in deposition due to existing 

regulations) and pre-emission (i.e., 

background) conditions. Second, to 

estimate the recreational fishing impacts 

of aquatic acidification in these lakes, 

an existing model of recreational fishing 

demand and site choice was applied. 

This model predicts how recreational 

fishing patterns in the Adirondacks 

would differ and how much higher the 

average annual value of recreational 

fishing services would be for New York 

residents if lake ANC levels 

corresponded to background (rather 

than business as usual) conditions. To 

estimate impacts on a broader category 

of cultural (and some provisioning) 

ecosystem services, results from the 

Banzhaf et al (2006) valuation survey of 

New York residents were adapted and 

applied to this context. The survey used 

a contingent valuation approach to 

estimate the average annual household 

willingness to pay (WTP) for future 

reductions in the percent of Adirondack 

lakes impaired by acidification. The 

focus of the survey was on impacts on 

aquatic resources. Pretesting of the 

survey indicated that respondents 

nonetheless tended to assume that 

benefits would occur in the condition of 

birds and forests as well as in 

recreational fishing. 

By extrapolating the 44 lake 

Adirondack case study to all 3,000 

Adirondack lakes and by applying the 

WTP survey results to all New York 

residents, the study estimated 

aggregated benefits between $300 and 

$800 million annually for the equivalent 

of improving lakes in the Adirondacks 

region to an ANC level of 50 μeq/L. The 

REA estimated 44 percent of the 

Adirondack lakes currently fall below 

an ANC of 50 μeq/L. Several states have 

set goals for improving the acid status 

of lakes and streams, generally targeting 

ANC in the range of 50 to 60 μeq/L, and 

have engaged in costly activities to 

decrease acidification. 

These results imply significant value 

to the public in addition to those 

derived from recreational fishing 

services. Note that the results are only 

applicable to improvements in the 

Adirondacks valued by residents of New 

York. If similar benefits exist in other 

acid-impacted areas, benefits for the 

nation as a whole could be substantial. 

The analysis provides results on only a 

subset of the impacts of acidification on 

ecosystem services and suggests that the 


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overall impact on these services is likely 

to be substantial. 

b. Terrestrial Acidification 

Chapters 4.4.3 and 4.4.4 of the PA 

review several economic studies of areas 

sensitive to terrestrial acidification. 

Forests in the northeastern U.S. provide 

several important and valuable 

provisioning ecosystem services, which 

are reflected in the production and sales 

of tree products. Sugar maples are a 

particularly important commercial 

hardwood tree species in the United 

States, producing timber and maple 

syrup that provide hundreds of millions 

of dollars in economic value annually. 

Red spruce is also used in a variety of 

wood products and provides up to $100 

million in economic value annually. 

Although the data do not exist to 

directly link acidification damages to 

economic values of lost recreational 

ecosystem services in forests, these 

resources are valuable to the public. A 

recent study, reviewed in the PA, 

suggests that the total annual value of 

recreational off-road driving was more 

than $9 billion and the value of hunting 

and wildlife viewing was more than $4 

billion each in the northeastern States. 

The EPA is not able to quantify at this 

time the specific effects on these values 

of acid deposition, or of any specific 

reductions in deposition, relative to the 

effects of many other factors that may 

affect them. 

c. Nutrient Enrichment 

Chapters 4.4.5 and 4.4.6 of the PA 

summarize economic studies of east 

coast estuaries affected by nutrient overenrichment or eutrophication. Estuaries 

in the eastern United States are 

important for fish and shellfish 

production. The estuaries are capable of 

supporting large stocks of resident 

commercial species, and they serve as 

the breeding grounds and interim 

habitat for several migratory species. To 

provide an indication of the magnitude 

of provisioning services associated with 

coastal fisheries, from 2005 to 2007, the 

average value of total catch was $1.5 

billion per year in 15 East Coast states. 

Estuaries also provide an important and 

substantial variety of cultural ecosystem 

services, including water-based 

recreational and aesthetic services. For 

example, data indicate that 4.8 percent 

of the population in coastal states from 

North Carolina to Massachusetts 

participated in saltwater fishing, with a 

total of 26 million saltwater fishing days 

in 2006. Based on estimates in the PA, 

total recreational value from these 

saltwater fishing days was 

approximately $1.3 billion. Recreational 

participation estimates for 1999–2000 

showed almost 6 million individuals 

participated in motorboating in coastal 

states from North Carolina to 

Massachusetts. The aggregate value of 

these coastal motorboating outings was 

$2 billion per year. EPA is not able to 

quantify at this time the specific effects 

on these values of nitrogen deposition, 

or of any specific reductions in 

deposition, relative to the effects of 

many other factors that may affect them. 

Terrestrial ecosystems can also suffer 

from nutrient over-enrichment. Each 

ecosystem is different in its composition 

of species and nutrient requirements. 

Changes to individual ecosystems from 

changes in nitrogen deposition can be 

hard to assess economically. Relative 

recreational values are often determined 

by public use information. Chapter 4.4.7 

of the PA reviewed studies related to 

park use in California. Data from 

California State Parks indicate that in 

2002, 68.7 percent of adult residents 

participated in trail hiking for an 

average of 24.1 days per year. The 

analyses in the PA indicate that the 

aggregate annual benefit for California 

residents from trail hiking in 2007 was 

$11.59 billion. EPA is not able to 

quantify at this time the specific effects 

on these values of nitrogen deposition, 

or of any specific reductions in 

deposition, relative to the effects of 

many other factors that may affect them. 

The PA also identified fire regulation 

as a service that could be affected by 

nutrient over-enrichment of the CSS and 

MCF ecosystems by encouraging growth 

of more flammable grasses, increasing 

fuel loads, and altering the fire cycle. 

Over the 5-year period from 2004 to 

2008, Southern California experienced, 

on average, over 4,000 fires per year, 

burning, on average, over 400,000 acres 

per year. It is not possible at this time 

to quantify the contribution of nitrogen 

deposition, among many other factors, 

to increased fire risk. 

3. Summary 

Adversity to public welfare can be 

understood by looking at how 

deposition of oxides of nitrogen and 

sulfur affect the ecological functions of 

an ecosystem (see II.A.), and then 

understanding the ecosystem services 

that are degraded. The monetized value 

of the ecosystem services provided by 

ecosystems that are sensitive to 

deposition of oxides of nitrogen and 

sulfur are in the billions of dollars each 

year, though it is not possible to 

quantify or monetize at this time the 

effects on these values of nitrogen and 

sulfur deposition or of any changes in 

deposition that may result from new 

secondary standards. Many lakes and 

streams are known to be degraded by 

acidic deposition which affects 

recreational fishing and tourism. Forest 

growth is likely suffering from acidic 

deposition in sensitive areas affecting 

red spruce and sugar maple timber 

production, sugar maple syrup 

production, hiking, aesthetic enjoyment 

and tourism. Nitrogen deposition 

contributes significantly to 

eutrophication in many estuaries 

affecting fish production, swimming, 

boating, aesthetic enjoyment and 

tourism. Ecosystem services are likely 

affected by nutrient enrichment in many 

natural and scenic terrestrial areas, 

affecting biodiversity, including habitat 

for rare and endangered species, fire 

control, hiking, aesthetic enjoyment and 

tourism. 

D. Adequacy of the Current Standards 

An important issue to be addressed in 

the current review of the secondary 

standards for oxides of nitrogen and 

sulfur is whether, in view of the 

scientific evidence reflected in the ISA, 

additional information on exposure and 

risk discussed in the REA, and 

conclusions drawn from the PA, the 

existing standards provide adequate 

protection. The Administrator therefore, 

has considered the extent to which the 

current standards are adequate for the 

protection of public welfare. Having 

reached the general conclusion that 

aquatic and terrestrial ecosystems can 

be degraded by deposition of oxides of 

nitrogen and sulfur, it is then necessary 

to first evaluate the appropriateness (in 

terms of form and structure) of the 

current standards to address the 

ecological effects of oxides of nitrogen 

and sulfur as well as the adequacy of the 

current secondary standards for oxides 

of nitrogen and sulfur to provide 

requisite protection by considering to 

what degree risks to sensitive 

ecosystems would be expected to occur 

in areas that meet the current standards. 

Conclusions regarding the adequacy of 

the current standards are based on the 

available ecological effects, exposure 

and risk-based evidence. In evaluating 

the strength of this information, EPA 

has taken into account the uncertainties 

and limitations in the scientific 

evidence. This section addresses the 

adequacy of the current standards to 

protect against direct exposure effects 

on plants from oxides of nitrogen and 

sulfur, the appropriateness of the 

current structure of the standards to 

address deposition-related effects of 

oxides of nitrogen and sulfur on 

sensitive ecosystems and finally, the 

adequacy of such standards to protect 

against adverse effects related to the 

deposition of oxides of nitrogen and 

sulfur. 


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3The current primary NO2 standard has recently 

been changed to the 3-year average of the 98th 

percentile of the annual distribution of the 1 hour 

daily maximum of the concentration of NO2. The 

current secondary standard remains as it was set in 

1971. 

1. Adequacy of the Current Standards 

for Direct Effects 

The current secondary oxides of 

nitrogen and sulfur standards are 

intended to protect against adverse 

effects to public welfare. For oxides of 

nitrogen, the current secondary standard 

was set identical to the primary 

standard,3 i.e., an annual standard set for 

NO2 to protect against adverse effects on 

vegetation from direct exposure to 

ambient oxides of nitrogen. For oxides 

of sulfur, the current secondary 

standard is a 3-hour standard intended 

to provide protection for plants from the 

direct foliar damage associated with 

atmospheric concentrations of SO2. It is 

appropriate to consider whether the 

current standards are adequate to 

protect against the direct effects on 

vegetation resulting from ambient NO2

and SO2 which were the basis for the 

current secondary standards. The ISA 

concluded that there was sufficient 

evidence to infer a causal relationship 

between exposure to SO2, NO, NO2 and 

PAN and injury to vegetation. 

Additional research on acute foliar 

injury has been limited and there is no 

evidence to suggest foliar injury below 

the levels of the current secondary 

standards for oxides of nitrogen and 

sulfur. There is sufficient evidence to 

suggest that the levels of the current 

standards are likely adequate to protect 

against direct phytotoxic effects. 

2. Appropriateness and Adequacy of the 

Current Standards for DepositionRelated Effects 

This section addresses two concepts 

necessary to evaluate the current 

standards in the context of deposition 

related effects. First, appropriateness of 

the current standards is considered with 

regard to indicator, form, level and 

averaging time. This discussion centers 

around the ability of the current 

standards to evaluate and provide 

protection against deposition related 

effects that vary spatially and 

temporally. It includes particular 

emphasis on the indicators and forms of 

the current standards and the degree to 

which they are ecologically relevant 

with regard to deposition related effects. 

Second, this section evaluates the 

current standards in terms of adequacy 

of protection. 

a. Appropriateness 

The ISA has established that the 

major effects of concern for this review 

of the oxides of nitrogen and sulfur 

standards are associated with deposition 

of nitrogen and sulfur caused by 

atmospheric concentrations of oxides of 

nitrogen and sulfur. The current 

standards are not directed toward 

depositional effects, and none of the 

elements of the current NAAQS— 

indicator, form, averaging time, and 

level—are suited for addressing the 

effects of nitrogen and sulfur deposition. 

Five issues arise that call into 

question the ecological relevance of the 

structure of the current secondary 

standards for oxides of nitrogen and 

sulfur. 

(1) The current SO2 secondary 

standard (0.5 ppm SO2 over a 3-hour 

average) does not utilize an exposure 

period that is relevant for ecosystem 

impacts. The majority of deposition 

related impacts are associated with 

depositional loads that occur over 

periods of months to years. This differs 

significantly from exposures associated 

with hourly concentrations of SO2 as 

measured by the current secondary 

standard. By addressing short-term 

concentrations, the current SO2

secondary standard, while protective 

against direct foliar effects from gaseous 

oxides of sulfur, does not take into 

account the findings of effects in the 

ISA, which notes the relationship 

between annual deposition of sulfur and 

acidification effects which are likely to 

be more severe and widespread than 

phytotoxic effects under current 

ambient conditions, and include effects 

from long term deposition as well as 

short term. Acidification is a process 

that occurs over time because the ability 

of an aquatic system to counteract acidic 

inputs is reduced as natural buffers are 

used more rapidly than they can be 

replaced through geologic weathering. 

The relevant period of exposure for 

ecosystems is, therefore, not the 

exposures captured in the short 

averaging time of the current SO2

secondary standard. The current 

secondary standard for oxides of 

nitrogen is an annual standard (0.053 

ppm averaged over 1 year) and as such 

is more ecologically relevant. 

(2) Current standards do not utilize 

appropriate atmospheric indicators. 

Nitrogen dioxide and SO2 are used as 

the component of oxides of nitrogen and 

sulfur that are measured, but they do 

not provide a complete link to the direct 

effects on ecosystems from deposition of 

oxides of nitrogen and sulfur as they do 

not capture all relevant chemical 

species of oxidized nitrogen and 

oxidized sulfur that contribute to 

deposition. The ISA provides evidence 

that deposition related effects are linked 

with total nitrogen and total sulfur 

deposition, and thus all forms of 

oxidized nitrogen and oxidized sulfur 

that are deposited will contribute to 

effects on ecosystems. Thus, by using 

atmospheric NO2 and SO2

concentrations as indicators, the current 

standards address only a fraction of total 

atmospheric oxides of nitrogen and 

sulfur, and do not take into account the 

effects from deposition of total 

atmospheric oxides of nitrogen and 

sulfur. This suggests that more 

comprehensive atmospheric indicators 

should be considered in designing 

ecologically relevant standards. 

(3) Current standards reflect separate 

assessments of the two individual 

pollutants, NO2 and SO2, rather than 

assessing the joint impacts of deposition 

to ecosystems. Recognizing the role that 

each pollutant plays in jointly affecting 

ecosystem indicators, functions, and 

services is vital to developing a 

meaningful standard. The clearest 

example of this interaction is in 

assessment of the impacts of acidifying 

deposition on aquatic ecosystems. 

Acidification in an aquatic ecosystem 

depends on the total acidifying potential 

of the deposition of both nitrogen and 

sulfur from both atmospheric deposition 

of oxides of nitrogen and sulfur as well 

as the inputs from other sources of 

nitrogen and sulfur such as reduced 

nitrogen and non-atmospheric sources. 

It is the joint impact of the two 

pollutants that determines the ultimate 

effect on organisms within the 

ecosystem, and critical ecosystem 

functions such as habitat provision and 

biodiversity. Standards that are set 

independently are less able to account 

for the contribution of the other 

pollutant. This suggests that interactions 

between oxides of nitrogen and oxides 

of sulfur should be a critical element of 

the conceptual framework for 

ecologically relevant standards. There 

are also important interactions between 

oxides of nitrogen and sulfur and 

reduced forms of nitrogen, which also 

contribute to acidification and nutrient 

enrichment. It is important that the 

structure of the standards address the 

role of reduced nitrogen in determining 

the ecological effects resulting from 

deposition of atmospheric oxides of 

nitrogen and sulfur. Consideration will 

also have to be given to total loadings 

as ecosystems respond to all sources of 

nitrogen and sulfur. 

(4) Current standards do not take into 

account variability in ecosystem 

sensitivity. Ecosystems are not 

uniformly distributed either spatially or 

temporally in their sensitivity to oxides 

of nitrogen and sulfur. Therefore, failure 

to account for the major determinants of 

variability, including geological and soil 


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characteristics related to the sensitivity 

to acidification or nutrient enrichment 

as well as atmospheric and landscape 

characteristics that govern rates of 

deposition, may lead to standards that 

do not provide requisite levels of 

protection across ecosystems. The 

current structures of the standards do 

not address the complexities in the 

responses of ecosystems to deposition of 

oxides of nitrogen and sulfur. 

Ecosystems contain complex groupings 

of organisms that respond in various 

ways to the alterations of soil and water 

that result from deposition of nitrogen 

and sulfur compounds. Different 

ecosystems therefore respond in 

different ways depending on a 

multitude of factors that control how 

deposition is integrated into the system. 

For example, the same levels of 

deposition falling on limestone 

dominated soils have a very different 

effect from those falling on shallow 

glaciated soils underlain with granite. 

One system may over time display no 

obvious detriment while the other may 

experience a catastrophic loss in fish 

communities. This degree of sensitivity 

is a function of many atmospheric 

factors that control rates of deposition as 

well as ecological factors that control 

how an ecosystem responds to that 

deposition. The current standards do 

not take into account spatial and 

seasonal variations, not only in 

depositional loadings, but also in 

sensitivity of ecosystems exposed to 

those loadings. Based on the discussion 

summarized above, the PA concludes 

that the current secondary standards for 

oxides of nitrogen and oxides of sulfur 

are not ecologically relevant in terms of 

averaging time, form, level or indicator. 

b. Adequacy of Protection 

As described in the PA, ambient 

conditions in 2005 indicate that the 

current SO2 and NO2 secondary 

standards were not exceeded at that 

time (US EPA, 2011, Figures 6–1 and 6– 

2) in locations where negative ecological 

effects have been observed. In many 

locations, SO2 and NO2 concentrations 

are substantially below the levels of the 

secondary standards. This pattern 

suggests that levels of deposition and 

any negative effects on ecosystems due 

to deposition of oxides of nitrogen and 

sulfur under recent conditions are 

occurring even though areas meet or are 

below current standards. In addition, 

based on conclusions in the REA, these 

levels will not decline in the future to 

levels below which it is reasonable to 

anticipate effects. 

In determining the adequacy of the 

current secondary standards for oxides 

of nitrogen and sulfur the PA 

considered the extent to which ambient 

deposition contributes to loadings in 

ecosystems. Since the last review of the 

secondary standard for oxides of 

nitrogen, a great deal of information on 

the contribution of atmospheric 

deposition associated with ambient 

oxides of nitrogen has become available. 

The REA presents a thorough 

assessment of the contribution of 

oxidized nitrogen to nitrogen deposition 

throughout the U.S., and the relative 

contributions of ambient oxidized and 

reduced forms of nitrogen. The REA 

concludes that based on that analysis, 

ambient oxides of nitrogen are a 

significant component of atmospheric 

nitrogen deposition, even in areas with 

relatively high rates of deposition of 

reduced nitrogen. In addition, 

atmospheric deposition of oxidized 

nitrogen contributes significantly to 

total nitrogen loadings in nitrogen 

sensitive ecosystems. 

The ISA summarizes the available 

studies of relative nitrogen contribution 

and finds that in much of the U.S., 

oxides of nitrogen contribute from 50 to 

75 percent of total atmospheric 

deposition relative to total reactive 

nitrogen, which includes oxidized and 

reduced nitrogen species (US EPA, 

2008, section 2.8.4). Although the 

proportion of total nitrogen loadings 

associated with atmospheric deposition 

of nitrogen varies across locations, the 

ISA indicates that atmospheric nitrogen 

deposition is the main source of new 

anthropogenic nitrogen to most 

headwater streams, high elevation lakes, 

and low-order streams. Atmospheric 

nitrogen deposition contributes to the 

total nitrogen load in terrestrial, 

wetland, freshwater and estuarine 

ecosystems that receive nitrogen 

through multiple pathways. In several 

large estuarine systems, including the 

Chesapeake Bay, atmospheric 

deposition accounts for between 10 and 

40 percent of total nitrogen loadings (US 

EPA, 2008). 

Atmospheric concentrations of oxides 

of sulfur account for nearly all sulfur 

deposition in the US. For the period 

2004–2006, mean sulfur deposition in 

the U.S. was greatest east of the 

Mississippi River with the highest 

deposition amount, 21.3 kg S/ha-yr, in 

the Ohio River Valley where most 

recording stations reported 3-year 

averages >10 kg S/ha-yr. Numerous 

other stations in the East reported S 

deposition >5 kg S/ha-yr. Total sulfur 

deposition in the U.S. west of the 100th 

meridian was relatively low, with all 

recording stations reporting <2 kg S/hayr and many reporting <1 kg S/ha-yr. 

Sulfur was primarily deposited in the 

form of wet SO4

2¥ followed in 

decreasing order by a smaller proportion 

of dry SO2 and a much smaller 

proportion of deposition as dry SO4

2¥. 

As discussed throughout the REA (US 

EPA, 2009 and section II.B above), there 

are several key areas of risk that are 

associated with ambient concentrations 

of oxides of nitrogen and sulfur. As 

noted earlier, in previous reviews of the 

secondary standards for oxides of 

nitrogen and sulfur, the standards were 

designed to protect against direct 

exposure of plants to ambient 

concentrations of the pollutants. A 

significant shift in understanding of the 

effects of oxides of nitrogen and sulfur 

has occurred since the last reviews, 

reflecting the large amount of research 

that has been conducted on the effects 

of deposition of nitrogen and sulfur to 

ecosystems. The most significant 

current risks of adverse effects to public 

welfare are those related to deposition 

of oxides of nitrogen and sulfur to both 

terrestrial and aquatic ecosystems. 

These risks fall into two categories, 

acidification and nutrient enrichment, 

which were emphasized in the REA as 

most relevant to evaluating the 

adequacy of the existing standards in 

protecting public welfare from adverse 

ecological effects. 

i. Aquatic Acidification 

The focus of the REA case studies was 

on determining whether deposition of 

sulfur and oxidized nitrogen in 

locations where ambient oxides of 

nitrogen and sulfur were at or below the 

current standards was resulting in 

acidification and related effects, 

including episodic acidification and 

mercury methylation. Based on the case 

studies conducted for lakes in the 

Adirondacks and streams in 

Shenandoah National Park (case studies 

are discussed more fully in section II.B 

and US EPA, 2009), there is significant 

risk to acid sensitive aquatic ecosystems 

at atmospheric concentrations of oxides 

of nitrogen and sulfur at or below the 

current standards. The REA also 

supports strongly a relationship 

between atmospheric deposition of 

oxides of nitrogen and sulfur and loss of 

ANC in sensitive ecosystems and 

indicates that ANC is an excellent 

indicator of aquatic acidification. The 

REA also concludes that at levels of 

deposition associated with oxides of 

nitrogen and sulfur concentrations at or 

below the current standards, ANC levels 

are expected to be below benchmark 

values that are associated with 

significant losses in fish species 

richness. 

Significant portions of the U.S. are 

acid sensitive, and current deposition 

levels exceed those that would allow 


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recovery of the most acid sensitive lakes 

in the Adirondacks (US EPA, 2008, 

Executive Summary). In addition, 

because of past loadings, areas of the 

Shenandoah are sensitive to current 

deposition levels (US EPA, 2008, 

Executive Summary). Parts of the West 

are naturally less sensitive to 

acidification and subjected to lower 

deposition (particularly SOX) levels 

relative to the eastern United States, and 

as such, less focus in the ISA is placed 

on the adequacy of the existing 

standards in these areas, with the 

exception of the mountainous areas of 

the West, which experience episodic 

acidification due to deposition. 

In describing the effects of 

acidification in the two case study areas 

the REA uses the approach of describing 

benchmarks in terms of ANC values. 

Many locations in sensitive areas of the 

U.S. have ANC levels below benchmark 

levels for ANC classified as severe, 

elevated, or moderate concern (US EPA, 

2011, Figure 2–1). The average current 

ANC levels across 44 lakes in the 

Adirondack case study area is 62.1 μeq/ 

L (moderate concern). However, 44 

percent of lakes had deposition levels 

exceeding the critical load for an ANC 

of 50 μeq/L (elevated), and 28 percent of 

lakes had deposition levels exceeding 

the (higher) critical load for an ANC of 

20 μeq/L (severe) (US EPA, 2009, 

section 4.2.4.2). This information 

indicates that almost half of the 44 lakes 

in the Adirondacks case study area are 

at an elevated concern level, and almost 

a third are at a severe concern level. 

These levels are associated with greatly 

diminished fish species diversity, and 

losses in the health and reproductive 

capacity of remaining populations. 

Based on assessments of the 

relationship between number of fish 

species and ANC level in both the 

Adirondacks and Shenandoah areas, the 

number of fish species is decreased by 

over half at an ANC level of 20 μeq/L 

relative to an ANC level at 100 μeq/L 

(US EPA, 2009, Figure 4.2–1). When 

extrapolated to the full population of 

lakes in the Adirondacks area using 

weights based on the EMAP probability 

survey (US EPA, 2009, section 4.2.6.1), 

36 percent of lakes exceeded the critical 

load for an ANC of 50 μeq/L and 13 

percent of lakes exceeded the critical 

load for an ANC of 20 μeq/L. 

Many streams in the Shenandoah case 

study area also have levels of deposition 

that are associated with ANC levels 

classified as severe, elevated, or 

moderate concern. The average ANC 

under recent conditions for the 60 

streams evaluated in the Shenandoah 

case study area is 57.9 μeq/L, indicating 

moderate concern. However, 85 percent 

of these streams had recent deposition 

exceeding the critical load for an ANC 

of 50 μeq/L, and 72 percent exceeded 

the critical load for an ANC of 20 μeq/ 

L. As with the Adirondacks area, this 

information suggests that ANC levels 

may decline in the future and 

significant numbers of sensitive streams 

in the Shenandoah area are at risk of 

adverse impacts on fish populations if 

recent conditions persist. Many other 

streams in the Shenandoah area are also 

likely to experience conditions of 

elevated to severe concern based on the 

prevalence in the area of bedrock 

geology associated with increased 

sensitivity to acidification suggesting 

that effects due to stream acidification 

could be widespread in the Shenandoah 

area (US EPA, 2009, section 4.2.6.2). 

In addition to these chronic 

acidification effects, the ISA notes that 

‘‘consideration of episodic acidification 

greatly increases the extent and degree 

of estimated effects for acidifying 

deposition on surface waters’’ (US EPA, 

2008, section 3.2.1.6). Some studies 

show that the number of lakes that 

could be classified as acid-impacted 

based on episodic acidification is 2 to 3 

times the number of lakes classified as 

acid-impacted based on chronic ANC. 

These episodic acidification events can 

have long term effects on fish 

populations (US EPA, 2008, section 

3.2.1.6). Under recent conditions, 

episodic acidification has been observed 

in locations in the eastern U.S. and in 

the mountainous western U.S. (US EPA, 

2008, section 3.2.1.6). 

The ISA, REA and PA all conclude 

that the current standards are not 

adequate to protect against the adverse 

impacts of aquatic acidification on 

sensitive ecosystems. A recent survey, 

as reported in the ISA, found sensitive 

streams in many locations in the U.S., 

including the Appalachian Mountains, 

the Coastal Plain, and the Mountainous 

West (US EPA, 2008, section 4.2.2.3). In 

these sensitive areas, between 1 and 6 

percent of stream kilometers are 

chronically acidified. The REA further 

concludes that both the Adirondack and 

Shenandoah case study areas are 

currently receiving deposition from 

ambient oxides of nitrogen and sulfur in 

excess of their ability to neutralize such 

inputs. In addition, based on the current 

emission scenarios, forecast modeling 

out to the year 2020 as well as 2050 

indicates a large number of streams in 

these areas will still be adversely 

impacted (section II.B). Based on these 

considerations, the PA concludes that 

the current secondary NAAQS for 

oxides of nitrogen and sulfur do not 

provide adequate protection of sensitive 

ecosystems with regard to aquatic 

acidification. 

ii. Terrestrial Acidification 

Based on the terrestrial acidification 

case studies, Kane Experimental Forest 

in Pennsylvania and Hubbard Brook 

Experimental Forest described in 

section II.B) of sugar maple and red 

spruce habitat, the REA concludes that 

there is significant risk to sensitive 

terrestrial ecosystems from acidification 

at atmospheric concentrations of NO2

and SO2 at or below the current 

standards. The ecological indicator 

selected for terrestrial acidification is 

the BC/Al, which has been linked to tree 

health and growth. The results of the 

REA strongly support a relationship 

between atmospheric deposition of 

oxides of nitrogen and sulfur and BC/Al, 

and that BC/Al is a good indicator of 

terrestrial acidification. At levels of 

deposition associated with oxides of 

nitrogen and sulfur concentrations at or 

below the current standards, BC/Al 

levels are expected to be below 

benchmark values that are associated 

with significant effects on tree health 

and growth. Such degradation of 

terrestrial ecosystems could affect 

ecosystem services such as habitat 

provisioning, endangered species, goods 

production (timber, syrup, etc.) among 

others. 

Many locations in sensitive areas of 

the U.S. have BC/Al levels below 

benchmark levels classified as providing 

low to intermediate levels of protection 

to tree health. At a BC/Al ratio of 1.2 

(intermediate level of protection), red 

spruce growth can be reduced by 20 

percent. At a BC/Al ratio of 0.6 (low 

level of protection), sugar maple growth 

can be decreased by 20 percent. The 

REA did not evaluate broad sensitive 

regions. However, in the sugar maple 

case study area (Kane Experimental 

Forest), recent deposition levels are 

associated with a BC/Al ratio below 1.2, 

indicating between intermediate and 

low level of protection, which would 

indicate the potential for a greater than 

20 percent reduction in growth. In the 

red spruce case study area (Hubbard 

Brook Experimental Forest), recent 

deposition levels are associated with a 

BC/Al ratio slightly above 1.2, 

indicating slightly better than an 

intermediate level of protection (US 

EPA, 2009, section 4.3.5.1). 

Over the full range of sugar maple, 12 

percent of evaluated forest plots 

exceeded the critical loads for a BC/Al 

ratio of 1.2, and 3 percent exceeded the 

critical load for a BC/Al ratio of 0.6. 

However, there was large variability 

across states. In New Jersey, 67 percent 

of plots exceeded the critical load for a 


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BC/Al ratio of 1.2, while in several 

states on the outskirts of the range for 

sugar maple (e.g. Arkansas, Illinois) no 

plots exceeded the critical load for a BC/ 

Al ratio of 1.2. For red spruce, overall 

5 percent of plots exceeded the critical 

load for a BC/Al ratio of 1.2, and 3 

percent exceeded the critical load for a 

BC/Al ratio of 0.6. In the major red 

spruce producing states (Maine, New 

Hampshire, and Vermont), critical loads 

for a BC/Al ratio of 1.2 were exceeded 

in 0.5, 38, and 6 percent of plots, 

respectively. 

The ISA, REA and PA all conclude 

that the current standards are not 

adequate to protect against the adverse 

impacts of terrestrial acidification on 

sensitive ecosystems. As stated in the 

REA and PA, the main drawback, with 

the understanding of terrestrial 

acidification lies in the sparseness of 

available data by which we can predict 

critical loads and that the data are based 

on laboratory responses rather than field 

measurements. Other stressors that are 

present in the field but that are not 

present in the laboratory may confound 

this relationship. The REA does 

however, conclude that the case study 

results, when extended to a 27 state 

region, show that nitrogen and sulfur 

acidifying deposition in the sugar maple 

and red spruce forest areas caused the 

calculated Bc/Al ratio to fall below 1.2 

(the intermediate level of protection) in 

12 percent of the sugar maple plots and 

5 percent of the red spruce plots; 

however, results from individual states 

ranged from 0 to 67 percent of the plots 

for sugar maple and 0 to 100 percent of 

the plots for red spruce. 

iii. Terrestrial Nutrient Enrichment 

Nutrient enrichment effects are due to 

nitrogen loadings from both 

atmospheric and non-atmospheric 

sources. Evaluation of nutrient 

enrichment effects requires an 

understanding that nutrient inputs are 

essential to ecosystem health and that 

specific long term levels of nutrients in 

a system affect the types of species that 

occur over long periods of time. Short 

term additions of nutrients can affect 

species competition, and even small 

additions of nitrogen in areas that are 

traditionally nutrient poor can have 

significant impacts on productivity as 

well as species composition. Most 

ecosystems in the U.S. are nitrogenlimited, so regional decreases in 

emissions and deposition of airborne 

nitrogen compounds could lead to some 

decrease in growth of the vegetation that 

surrounds the targeted aquatic system 

but as discussed below evidence for this 

is mixed. Whether these changes in 

plant growth are seen as beneficial or 

adverse will depend on the nature of the 

ecosystem being assessed. 

Information on the effects of changes 

in nitrogen deposition on forestlands 

and other terrestrial ecosystems is very 

limited. The multiplicity of factors 

affecting forests, including other 

potential stressors such as ozone, and 

limiting factors such as moisture and 

other nutrients, confound assessments 

of marginal changes in any one stressor 

or nutrient in forest ecosystems. The 

ISA notes that only a fraction of the 

deposited nitrogen is taken up by the 

forests, most of the nitrogen is retained 

in the soils (US EPA, 2008, section 

3.3.2.1). In addition, the ISA indicates 

that forest management practices can 

significantly affect the nitrogen cycling 

within a forest ecosystem, and as such, 

the response of managed forests to 

nitrogen deposition will be variable 

depending on the forest management 

practices employed in a given forest 

ecosystem (US EPA, 2008, Annex C 

C.6.3). Increases in the availability of 

nitrogen in nitrogen-limited forests via 

atmospheric deposition could increase 

forest production over large nonmanaged areas, but the evidence is 

mixed, with some studies showing 

increased production and other showing 

little effect on wood production (US 

EPA, 2008, section 3.3.9). Because 

leaching of nitrate can promote cation 

losses, which in some cases create 

nutrient imbalances, slower growth and 

lessened disease and freezing tolerances 

for forest trees, the net effect of 

increased N on forests in the U.S. is 

uncertain (US EPA, 2008, section 3.3.9). 

The scientific literature has many 

examples of the deleterious effects 

caused by excessive nitrogen loadings to 

terrestrial systems. Several studies have 

set benchmark values for levels of N 

deposition at which scientifically 

adverse effects are known to occur. 

Large areas of the country appear to be 

experiencing deposition above these 

benchmarks. The ISA indicates studies 

that have found that at 3.1 kg N/ha/yr, 

the community of lichens begins to 

change from acidophytic to tolerant 

species; at 5.2 kg N/ha/yr, the typical 

dominance by acidophytic species no 

longer occurs; and at 10.2 kg N/ha/yr, 

acidophytic lichens are totally lost from 

the community. Additional studies in 

the Colorado Front Range of the Rocky 

Mountain National Park support these 

findings. These three values (3.1, 5.2, 

and 10.2 kg/ha/yr) are one set of 

ecologically meaningful benchmarks for 

the mixed conifer forest (MCF) of the 

pacific coast regions. Nearly all of the 

known sensitive communities receive 

total nitrogen deposition levels above 

the 3.1 N kg/ha/yr ecological benchmark 

according to the 12 km, 2002 CMAQ/ 

NADP data, with the exception of the 

easternmost Sierra Nevadas. The MCFs 

in the southern portion of the Sierra 

Nevada forests and nearly all MCF 

communities in the San Bernardino 

forests receive total nitrogen deposition 

levels above the 5.2 N kg/ha/yr 

ecological benchmark. 

Coastal Sage Scrub communities are 

also known to be sensitive to 

community shifts caused by excess 

nitrogen loadings. Studies have 

investigated the amount of nitrogen 

utilized by healthy and degraded CSS 

systems. In healthy stands, the authors 

estimated that 3.3 kg N/ha/yr was used 

for CSS plant growth. It is assumed that 

3.3 kg N/ha/yr is near the point where 

nitrogen is no longer limiting in the CSS 

community and above which level 

community changes occur, including 

dominance by invasive species and loss 

of coastal sage scrub. Therefore, this 

amount can be considered an ecological 

benchmark for the CSS community. The 

majority of the known CSS range is 

currently receiving deposition in excess 

of this benchmark. Thus, the REA 

concludes that recent conditions where 

oxides of nitrogen ambient 

concentrations are at or below the 

current oxides of nitrogen secondary 

standards are not adequate to protect 

against anticipated adverse impacts 

from N nutrient enrichment in sensitive 

ecosystems. 

iv. Aquatic Nutrient Enrichment 

The REA aquatic nutrient enrichment 

case studies focused on coastal estuaries 

and revealed that while current ambient 

loadings of atmospheric oxides of 

nitrogen are contributing to the overall 

depositional loading of coastal estuaries, 

other non-atmospheric sources are 

contributing in far greater amounts in 

total, although atmospheric 

contributions are as large as some other 

individual source types. The ability of 

current data and models to characterize 

the incremental adverse impacts of 

nitrogen deposition is limited, both by 

the available ecological indicators, and 

by the inability to attribute specific 

effects to atmospheric sources of 

nitrogen. The REA case studies used 

ASSETS EI as the ecological indicator 

for aquatic nutrient enrichment. This 

index is a six level index characterizing 

overall eutrophication risk in a water 

body. This indicator is not sensitive to 

changes in nitrogen deposition within a 

single level of the index. In addition, 

this type of indicator does not reflect the 

impact of nitrogen deposition in 

conjunction with other sources of 

nitrogen. 


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4The annual secondary standard for oxides of 

nitrogen is being specified in units of ppb to 

conform to the current version of the annual 

primary standard, as specified in the final rule for 

the most recent review of the NO2 primary NAAQS 

(75 FR 6531; February 9, 2010). 

Based on the above considerations, 

the REA concludes that the ASSETS EI 

is not an appropriate ecological 

indicator for estuarine aquatic 

eutrophication and that additional 

analysis is required to develop an 

appropriate indicator for determining 

the appropriate levels of protection from 

N nutrient enrichment effects in 

estuaries related to deposition of oxides 

of nitrogen. As a result, EPA is unable 

to make a determination as to the 

adequacy of the existing secondary 

oxides of nitrogen standard in 

protecting public welfare from nitrogen 

nutrient enrichment effects in estuarine 

aquatic ecosystems. 

Additionally, nitrogen deposition can 

alter species composition and cause 

eutrophication in freshwater systems. In 

the Rocky Mountains, for example, 

deposition loads of 1.5 to 2 kg/ha/yr 

which are well within current ambient 

levels are known to cause changes in 

species composition in diatom 

communities indicating impaired water 

quality (US EPA, 2008, section 3.3.5.3). 

This suggests that the existing 

secondary standard for oxides of 

nitrogen does not protect such 

ecosystems and their resulting services 

from impairment. 

v. Other Effects 

An important consideration in 

looking at the effects of deposition of 

oxides of sulfur in aquatic ecosystems is 

the potential for production of MeHg, a 

neurotoxic contaminant. The 

production of meaningful amounts of 

MeHg requires the presence of SO4

2¥

and mercury, and where mercury is 

present, increased availability of SO4

2¥

results in increased production of 

MeHg. There is increasing evidence on 

the relationship between sulfur 

deposition and increased methylation of 

mercury in aquatic environments; this 

effect occurs only where other factors 

are present at levels within a range to 

allow methylation. The production of 

MeHg requires the presence of SO4

2¥

and mercury, but the amount of MeHg 

produced varies with oxygen content, 

temperature, pH and supply of labile 

organic carbon (US EPA, 2008, section 

3.4). In watersheds where changes in 

sulfate deposition did not produce an 

effect, one or several of those interacting 

factors were not in the range required 

for meaningful methylation to occur (US 

EPA, 2008, section 3.4). Watersheds 

with conditions known to be conducive 

to mercury methylation can be found in 

the northeastern United States and 

southeastern Canada (US EPA, 2009, 

section 6). 

With respect to sulfur deposition and 

mercury methylation, the final ISA 

determined that ’’[t]he evidence is 

sufficient to infer a causal relationship 

between sulfur deposition and increased 

mercury methylation in wetlands and 

aquatic environments.’’ However, EPA 

did not conduct a quantitative 

assessment of the risks associated with 

increased mercury methylation under 

current conditions. As such, EPA is 

unable to make a determination as to the 

adequacy of the existing SO2 secondary 

standards in protecting against welfare 

effects associated with increased 

mercury methylation. 

vi. Summary of Adequacy 

Considerations 

In summary, the PA concludes that 

currently available scientific evidence 

and assessments clearly call into 

question the adequacy of the current 

standards with regard to depositionrelated effects on sensitive aquatic and 

terrestrial ecosystems, including 

acidification and nutrient enrichment. 

Further, the PA recognizes that the 

elements of the current standards— 

indicator, averaging time, level and 

form—are not ecologically relevant, and 

are thus not appropriate for standards 

designed to provide such protection. 

Thus, the PA concludes that 

consideration should be given to 

establishing a new ecologically relevant 

multi-pollutant, multimedia standard to 

provide appropriate protection from 

deposition-related ecological effects of 

oxides of nitrogen and sulfur on 

sensitive ecosystems, with a focus on 

protecting against adverse effects 

associated with acidifying deposition in 

sensitive aquatic ecosystems. 

3. CASAC Views 

In a letter to the Administrator 

(Russell and Samet 2011a), the CASAC 

Oxides of Nitrogen and Oxides of Sulfur 

Panel, with full endorsement of the 

chartered CASAC, unanimously 

concluded that: 

EPA staff has demonstrated through the 

Integrated Science Assessment (ISA), Risk 

and Exposure Characterization (REA) and the 

draft PA that ambient NOX and SOX can 

have, and are having, adverse environmental 

impacts. The Panel views that the current 

NOX and SOX secondary standards should be 

retained to protect against direct adverse 

impacts to vegetation from exposure to gas 

phase exposures of these two families of air 

pollutants. Further, the ISA, REA and draft 

PA demonstrate that adverse impacts to 

aquatic ecosystems are also occurring due to 

deposition of NOX and SOX. Those impacts 

include acidification and undesirable levels 

of nutrient enrichment in some aquatic 

ecosystems. The levels of the current NOX

and SOX secondary NAAQS are not 

sufficient, nor the forms of those standards 

appropriate, to protect against adverse 

depositional effects; thus a revised NAAQS is 

warranted. 

In addition, with regard to the joint 

consideration of both oxides of nitrogen 

and oxides of sulfur as well as the 

consideration of deposition related 

effects, CASAC concluded that the PA 

had developed a credible methodology 

for considering such effects. The Panel 

stated that ‘‘the Policy Assessment 

develops a framework for a multipollutant, multimedia standard that is 

ecologically relevant and reflects the 

combined impacts of these two 

pollutants as they deposit to sensitive 

aquatic ecosystems.’’ 

4. Administrator’s Proposed 

Conclusions Concerning Adequacy of 

Current Standard 

Based on the above considerations 

and taking into account CASAC advice, 

the Administrator recognizes that the 

purpose of the secondary standard is to 

protect against ‘‘adverse’’ effects 

resulting from exposure to oxides of 

nitrogen and sulfur, discussed above in 

section II.A. The Administrator also 

recognizes the need for conclusions as 

to the adequacy of the current standards 

for both direct and deposition related 

effects as well as conclusions as to the 

appropriateness and ecological 

relevance of the current standards. 

In considering what constitutes an 

ecological effect that is also adverse to 

the public welfare, the Administrator 

took into account the ISA conclusions 

regarding the nature and strength of the 

effects evidence, the risk and exposure 

assessment results, the degree to which 

the associated uncertainties should be 

considered in interpreting the results, 

the conclusions presented in the PA, 

and the views of CASAC and members 

of the public. On these bases, the 

Administrator concludes that the 

current secondary standards are 

adequate to protect against direct 

phytotoxic effects on vegetation. Thus, 

the Administrator proposes to retain the 

current secondary standard for oxides of 

nitrogen at 53 ppb,4 annual average 

concentration, measured in the ambient 

air as NO2, and the current secondary 

standard for oxides of sulfur at 0.5 ppm, 


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3-hour average concentration, measured 

in the ambient air as SO2. 

With regard to deposition-related 

effects, the Administrator has first to 

consider the appropriateness of the 

structure of the current standards to 

address ecological effects of concern. 

Based on the evidence as well as 

considering the advice given by CASAC 

on this matter, the Administrator 

concludes that the elements of the 

current standards are not ecologically 

relevant and thus are not appropriate to 

provide protection of ecosystems. On 

the subject of adequacy of protection 

with regard to deposition-related effects, 

the Administrator considered the full 

nature of ecological effects related to the 

deposition of ambient oxides of nitrogen 

and sulfur into sensitive ecosystems 

across the country. Her conclusions are 

based on the evidence presented in the 

ISA with regard to acidification and 

nutrient enrichment effects, the findings 

of the REA with regard to scope and 

severity of the current and likely future 

effects of deposition, the synthesis of 

both the scientific evidence and risk and 

exposure results in the PA as to the 

adequacy of the current standards, and 

the advice of both CASAC and the 

public. After such consideration, the 

Administrator concludes that current 

levels of oxides of nitrogen and sulfur 

are sufficient to cause acidification of 

both aquatic and terrestrial ecosystems, 

nutrient enrichment of terrestrial 

ecosystems and contribute to nutrient 

enrichment effects in estuaries that 

could be considered adverse, and the 

current secondary standards do not 

provide adequate protection from such 

effects. 

Having reached these conclusions, the 

Administrator determines that it is 

appropriate to consider alternative 

standards that are ecologically relevant. 

These considerations support the 

conclusion that the current secondary 

standards is neither appropriate nor 

adequate to protect against deposition 

related effects. The Administrator’s 

consideration of such alternative 

standards is discussed below in Section 

III. 

III. Rationale for Proposed Decision on 

Alternative Multi-Pollutant Approach 

to Secondary Standards for Aquatic 

Acidification 

Having reached the conclusion that 

the current NO2 and SO2 secondary 

standards are not adequate to provide 

appropriate protection against 

deposition-related effects associated 

with oxides of nitrogen and sulfur, the 

Administrator then considered what 

new multi-pollutant standard might be 

appropriate, at this time, to address 

such effects on public welfare. The 

Administrator recognizes that the 

inherently complex and variable 

linkages between ambient 

concentrations of nitrogen and sulfur 

oxides, the related deposited forms of 

nitrogen and sulfur, and the ecological 

responses that are associated with 

public welfare effects call for 

consideration of an ecologically relevant 

design of a standard that reflects these 

linkages. The Administrator also 

recognizes that characterization of such 

complex and variable linkages will 

necessarily require consideration of 

information and analyses that have 

important limitations and uncertainties. 

Despite its complexity, an 

ecologically relevant multi-pollutant 

standard to address deposition-related 

effects could still appropriately be 

defined in terms of the same basic 

elements that are used to define any 

NAAQS—indicator, form, averaging 

time, and level. The form would 

incorporate additional structural 

elements that reflect relevant multipollutant and multimedia attributes. 

These structural elements include the 

use of an ecological indicator, tied to the 

ecological effect we are focused on, and 

other elements that account for 

ecologically relevant factors other than 

ambient air concentrations. All of these 

elements would be needed to enable a 

linkage from ambient air indicators to 

the ecological indicator to define an 

ecologically relevant standard. As a 

result, such a standard would 

necessarily be more complex than the 

NAAQS that have been set historically 

to address effects associated with 

ambient concentrations of a single 

pollutant. 

More specifically, the Administrator 

considered an ecologically relevant 

multi-pollutant standard to address 

effects associated with acidifying 

deposition related to ambient 

concentrations of oxides of nitrogen and 

sulfur in sensitive aquatic ecosystems. 

This focus is consistent with the 

information presented in the ISA, REA, 

and PA, which highlighted the 

sufficiency of the quantity and quality 

of the available evidence and 

assessments associated with aquatic 

acidification relative to the information 

and assessments available for other 

deposition-related effects, including 

terrestrial acidification and aquatic and 

terrestrial nutrient enrichment. Based 

on its review of these documents, 

CASAC agreed that aquatic acidification 

should be the focus for developing a 

new multi-pollutant standard in this 

review. In reaching conclusions about 

an air quality standard designed to 

address deposition-related aquatic 

acidification effects, the Administrator 

also recognizes that such a standard 

may also provide some degree of 

protection against other depositionrelated effects. 

As discussed in chapter 7 of the PA, 

the development of a new multipollutant standard to address 

deposition-related aquatic acidification 

effects recognizes the need for 

consideration of a nationally applicable 

standard for protection against adverse 

effects of aquatic acidification on public 

welfare, while recognizing the complex 

and heterogeneous interactions between 

ambient air concentrations of nitrogen 

and sulfur oxides, the related deposition 

of nitrogen and sulfur, and associated 

ecological responses. The development 

of such a standard also needs to take 

into account the limitations and 

uncertainties in the available 

information and analyses upon which 

characterization of such interactions are 

based. The approach used in the PA also 

recognizes that while such a standard 

would be national in scope and 

coverage, the effects to public welfare 

from aquatic acidification will not occur 

to the same extent in all locations in the 

U.S., given the inherent variability of 

the responses of aquatic systems to the 

effects of acidifying deposition. 

As discussed above in section II, 

many locations in the U.S. are naturally 

protected against acid deposition due to 

underlying geological conditions. 

Likewise, some locations in the U.S., 

including lands managed for 

commercial agriculture and forestry, are 

not likely to be negatively impacted by 

current levels of nitrogen and sulfur 

deposition. As a result, while a new 

ecologically relevant secondary 

standard would apply everywhere, it 

would be structured to account for 

differences in the sensitivity of 

ecosystems across the country. This 

would allow for appropriate protection 

of sensitive aquatic ecosystems, which 

are relatively pristine and wild and 

generally in rural areas, and the services 

provided by such sensitive ecosystems, 

without requiring more protection than 

is needed elsewhere. 

As discussed below, the multipollutant standard developed in the PA 

would employ (1) total reactive oxidized 

nitrogen (NOy) and SOX as the 

atmospheric ambient air indicators; (2) 

a form that takes into account variable 

factors, such as atmospheric and 

ecosystem conditions that modify the 

amounts of deposited nitrogen and 

sulfur; the distinction between oxidized 

and reduced forms of nitrogen; effects of 

deposited nitrogen and sulfur on aquatic 

ecosystems in terms of the ecological 

indicator ANC; and the 


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5As discussed in chapter 2 of the PA, SO2 and 

particulate SO4 are routinely measured in ambient 

air monitoring networks, although only the Clean 

Air Status and Trends Network (CASTNET) filter 

packs do not intentionally exclude particle size 

fractions. The CMAQ treatment of SOX is the simple 

addition of both species, which are treated 

explicitly in the model formulation. All particle 

size fractions are included in the CMAQ SOX

estimates. 

representativeness of water bodies 

within a defined spatial area; (3) a 

multi-year averaging time, and (4) a 

standard level defined in terms of a 

single, national target ANC value that, 

in the context of the above form, 

identifies the levels of concentrations of 

NOy and SOX in the ambient air that 

would meet the standard. The form of 

such a standard has been defined by an 

index, AAI, which reflects the 

relationship between ambient 

concentrations of NOy and SOX and 

aquatic acidification effects that result 

from nitrogen and sulfur deposition 

related to these ambient concentrations. 

In presenting the considerations 

associated with such an air quality 

standard to address deposition-related 

aquatic acidification effects, the 

following sections focus on each 

element of the standard, including 

indicator (section III.A), form (section 

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

level (section III.D). Alternative 

combinations of levels and forms are 

discussed in section III.E. 

Considerations related to important 

uncertainties inherent in such an 

approach are discussed in section III.F. 

Advice from CASAC on such a new 

standard is presented in section III.G. 

The Administrator’s proposed decisions 

on such a new standard are presented in 

section III.H. 

A. Ambient Air Indicators 

In considering alternative ambient air 

indicators, the PA primarily focuses on 

the important attribute of association. 

Association in a broad sense refers to 

how well an ambient air indicator 

relates to the ecological effects of 

interest by virtue of both the framework 

that links the ambient indicator and 

effects and the empirical evidence that 

quantifies the linkages. The PA also 

considers how measurable or 

quantifiable an indicator is to enable its 

use as an effective indicator of relevant 

ambient air concentrations. 

As discussed above in section II.C, the 

PA concludes that indicators other than 

NO2 and SO2 should be considered as 

the appropriate indicators of oxides of 

nitrogen and sulfur in the ambient air 

for protection against the acidification 

effects associated with deposition of the 

associated nitrogen and sulfur. This 

conclusion is based on the recognition 

that all forms of nitrogen and sulfur in 

the ambient air contribute to deposition 

and resulting acidification, and as such, 

NO2 and SO2 are incomplete indicators. 

In principle, the ambient indicators 

should represent the species that are 

associated with oxides of nitrogen and 

sulfur in the ambient air and can 

contribute acidifying deposition. This 

includes both the species of oxides of 

nitrogen and sulfur that are directly 

emitted as well as species transformed 

in the atmosphere from oxides of 

nitrogen and sulfur that retain the 

nitrogen and sulfur atoms from directly 

emitted oxides of nitrogen and sulfur. 

All of these compounds are associated 

with oxides of nitrogen and sulfur in the 

ambient air and can contribute to 

acidifying deposition. 

The PA focuses in particular on the 

various compounds with nitrogen or 

sulfur atoms that are associated with 

oxides of nitrogen and sulfur, because 

the acidifying potential is specific to 

nitrogen and sulfur, and not other atoms 

(e.g., H, C, O) whether derived from the 

original source of oxides of nitrogen and 

sulfur emissions or from atmospheric 

transformations. For example, the 

acidifying potential of each molecule of 

NO2, NO, HNO3 or PAN is identical, as 

is the potential for each molecule of SO2

or ion of particulate sulfate, p-SO4. Each 

atom of sulfur affords twice the 

acidifying potential of each atom of 

nitrogen. 

1. Oxides of Sulfur 

As discussed in the PA (US EPA, 

2011, section 7.1.1), oxides of sulfur 

include the gases sulfur monoxide (SO), 

SO2, sulfur trioxide (SO3), disulfur 

monoxide (S2O), and particulate-phase 

sulfur compounds (referred to as SO4) 

that result from gas-phase sulfur oxides 

interacting with particles. However, the 

sum of SO2 and SO4 does represent 

virtually the entire ambient air mass of 

sulfur that contributes to acidification. 

In addition to accounting for virtually 

all the potential for acidification from 

oxidized sulfur in the ambient air, there 

are reliable methods to monitor the 

concentrations of SO2 and particulate 

SO4. In addition, much of the data used 

to develop the technical basis for the 

standard developed in the PA is based 

on monitoring or modeling of these 

species.5 The PA concludes that the 

sum of SO2 and SO4, referred to as SOX, 

are appropriate ambient air indicators of 

oxides of sulfur because they represent 

virtually all of the acidification 

potential of ambient air oxides of sulfur 

and there are reliable methods suitable 

for measuring SO2 and SO4. 

2. Oxides of Nitrogen 

As discussed in the PA (US EPA, 

2011, section 7.1.2), NOy, as defined in 

chapter 2 of the PA, incorporates 

basically all of the oxidized nitrogen 

species that have acidifying potential 

and as such, NOy should be considered 

as an appropriate indicator for oxides of 

nitrogen. Total reactive oxidized 

nitrogen is an aggregate measure of NO 

and NO2 and all of the reactive oxidized 

products of NO and NO2. That is, NOy

is a group of nitrogen compounds in 

which all of the compounds are either 

an oxide of nitrogen or compounds in 

which the nitrogen atoms came from 

oxides of nitrogen. Total reactive 

oxidized nitrogen is especially relevant 

as an ambient indicator for acidification 

in that it both relates to the oxides of 

nitrogen in the ambient air and also 

represents the acidification potential of 

all oxidized nitrogen species in the 

ambient air, whether an oxide of 

nitrogen or derived from oxides of 

nitrogen. 

There are currently available reliable 

methods of measuring aggregate NOy. 

The term ‘‘aggregate’’ measure means 

that the NOy, as measured, is not based 

on measuring each individual species of 

NOy and calculating an NOy value by 

summing the individual species. Rather, 

as described in chapter 2 of the PA, 

current measurement techniques 

process all of the individual NOy

species to produce a single aggregate 

measure of all of the nitrogen atoms 

associated with any NOy species. 

Consequently, the NOy measurement 

effectively provides the sum of all 

individual species, but the identity of 

the individual species is lost. As 

discussed above, the accounting for the 

individual nitrogen atoms is an 

accounting of the ambient air 

acidification potential of oxides of 

nitrogen and their transformation 

products and therefore the most relevant 

ambient indicator for aquatic 

acidification effects associated with 

oxides of nitrogen. 

This loss of the information on 

individual species motivated 

consideration of alternative or more 

narrowly defined indicators for oxides 

of nitrogen in the PA. Consideration of 

a subset of NOy species was based on 

the following reasoning. First, the actual 

dry deposition of nitrogen is determined 

on an individual species basis by 

multiplying the species concentration 

times a species-specific deposition 

velocity and then summed to develop 

an estimate of total dry deposition. 

Consequently, the use of individual 

ambient species has the potential to be 

more consistent with the underlying 


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6The PA also notes that NOy is a useful 

measurement for model evaluation purposes, which 

is especially important, recognizing the unique role 

that CMAQ plays in the development of this 

standard, as described below in section III.B. 

science of deposition and, therefore, has 

the potential to allow for a more 

rigorous evaluation of dry deposition 

with specialized field studies. In 

addition, there has been a suggestion of 

focusing only on the most quickly 

depositing NOy species, such as HNO3, 

as contributions from other NOy species 

such as NO2 may be negligible. These 

alternative indicators are discussed 

below. 

The PA considers the relative merits 

of using each individual NOy species as 

part of a group of indicators. In so 

doing, it was first noted that dry 

deposition of NOy is treated as the sum 

of the deposition of each individual 

species in advanced process-based air 

quality models like CMAQ, as described 

in chapter 2 of the PA. Conceptually one 

could extend this process-based 

approach by using all NOy species 

individually as separate indicators for 

oxides of nitrogen and requiring, for 

example, measurements of each of the 

species, including the dominant species 

of HNO3, particulate nitrate (p-NO3), 

true NO2, NO and PAN. The potential 

attraction of using individual species 

would be the reliance on actual 

deposition velocities. This could have 

more physical meaning in comparison 

to a constructed model of aggregate 

deposition of NOy, which is difficult to 

evaluate with observations because of 

the assimilation of many species with 

disparate deposition behavior. The PA 

notes that the major drawback of using 

individual NOy species as the indicators 

is the lack of reliable measurement 

techniques, especially for PAN and NO2

in rural locations, which renders the use 

of virtually any individual NOy species, 

except for NO and perhaps p-NO3, as 

functionally inadequate from a 

measurement perspective. 

The PA next considered the relative 

merits of using a subset of NOy species 

as the indicators for oxides of nitrogen, 

as was discussed above for oxides of 

sulfur. To the extent that certain species 

provide relatively minor contributions 

to total NOy deposition, it may be 

appropriate to consider excluding them 

as part of the indicator. As discussed in 

chapter 2 of the PA, each nitrogen 

species within the array of NOy species 

has species-specific dry deposition 

velocities. For example, the deposition 

velocity of HNO3 is much greater than 

the velocity for NO2 and, consequently, 

for a similar ambient air concentration, 

HNO3 contributes more deposition of 

acidifying nitrogen relative to NO2. In 

transitioning from source-oriented 

urban locations to rural environments, 

the ratio of the concentrations of HNO3

and PAN to NO2 increases. 

Based on the reasoning that a larger 

fraction of the deposited NOy is 

accounted for by total nitrate (the sum 

of HNO3 and p-NO3), a surrogate for the 

more rapidly depositing fraction of NOy, 

combined with the availability of 

reliable total nitrate measurements 

through the CASTNET, the PA 

considered using total nitrate as the 

indicator for oxides of nitrogen (US 

EPA, 2011, appendix E). Nitrate would 

be expected to correlate well with total 

reactive oxidized nitrogen deposition 

relative to NOy (US EPA, 2011, chapter 

2) despite the inherent noise associated 

with variable contributions of low 

deposition velocity species (e.g., NO2) 

that may have relatively high ambient 

concentrations. However, modeling 

simulations suggest that NOy may be a 

more robust indicator, relative to HNO3, 

in terms of relating absolute changes in 

ambient air concentrations to changes in 

nitrogen deposition driven by changes 

in ambient concentrations of oxides of 

nitrogen (US EPA, 2011, Figure 2–32). 

Based on the above considerations, 

the PA concludes that NOy should be 

considered as the appropriate ambient 

indicator for oxides of nitrogen based on 

its direct relationship to oxides of 

nitrogen in the ambient air and its direct 

relationship to deposition associated 

with aquatic acidification. Because NOy

represents all of the potentially 

acidifying oxidized nitrogen species in 

the ambient air, it is appropriately 

associated with the deposition of 

potentially acidifying compounds 

associated with oxides of nitrogen in the 

ambient air. In addition, there are 

reliable methods available to measure 

NOy. Measurement of each individual 

species of NOy, or the measurement of 

only a subset of species of NOy, is less 

appropriate because there are not 

reliable measurements methods 

available to measure all of the 

individual species of NOy and a subset 

of species would fail to account for 

significant portions of the oxidized 

reactive nitrogen that relate to 

acidification.6

B. Form 

Based on the evidence of the aquatic 

acidification effects caused by the 

deposition of NOy and SOX, the PA (US 

EPA, 2011, section 7.2) presents the 

development of a new form that is 

ecologically relevant for addressing 

such effects. The conceptual design for 

the form of such a standard includes 

three main components: an ecological 

indicator, deposition metrics that relate 

to the ecological indicator, and a 

function that relates ambient air 

indicators to deposition metrics. 

Collectively, these three components 

link the ecological indicator to ambient 

air indicators, as illustrated above in 

Fig II–1. 

The simplified flow diagram in 

Figure II–1 compresses the various 

atmospheric, biological, and 

geochemical processes associated with 

acidifying deposition to aquatic 

ecosystems into a simplified conceptual 

picture. The ecological indicator (left 

box) is related to atmospheric 

deposition through biogeochemical 

ecosystem models (middle box), which 

associate a target deposition load to a 

target ecological indicator. Once a target 

deposition is established, associated 

allowable air concentrations are 

determined (right box) through the 

relationships between concentration 

and deposition that are embodied in air 

quality models such as CMAQ. The 

following discussion describes the 

development and rationale for each of 

these components, as well as the 

integration of these components into the 

full expression of the form of the 

standard using the concept of a national 

AAI that represents a target ANC level 

as a function of ambient air 

concentrations. Spatial aggregation 

issues associated with defining each of 

the terms of this index are also 

addressed below. 

The AAI is designed to be an 

ecologically relevant form of the 

standard that determines the levels of 

NOy and SOX in the ambient air that 

would achieve a target ANC limit for the 

U.S. The intent of the AAI is to weight 

atmospheric concentrations of oxides of 

nitrogen and sulfur by their propensity 

to contribute to acidification through 

deposition, given the fundamental 

acidifying potential of each pollutant, 

and to take into account the ecological 

factors that govern acid sensitivity in 

different ecosystems. The index also 

accounts for the contribution of reduced 

nitrogen to acidification. Thus, the AAI 

encompasses those attributes of specific 

relevance to protecting ecosystems from 

the acidifying potential of ambient air 

concentrations of NOy and SOX. 

1. Ecological Indicator 

In considering alternative ecological 

indicators, the PA again primarily 

focuses on the attribute of association. 

In the case of an ecological indicator for 

aquatic acidification, association refers 

to the relationship between the 

indicator and adverse effects as 

discussed in section II. Because of the 

conceptual structure of the form of an 


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AAI-based standard (Figure III–1), this 

particular ecological indicator must also 

link up in a meaningful and quantifiable 

manner with acidifying atmospheric 

deposition. In effect, the ecological 

indicator for aquatic acidification is the 

bridge between biological impairment 

and deposition of NOy and SOX. 

This section presents the rationale in 

the PA for selecting ANC as the 

appropriate ecological indicator for 

consideration. Recognizing that ANC is 

not itself the causative or toxic agent for 

adverse aquatic acidification effects, the 

rationale for using ANC as the relevant 

ecological indicator is based on the 

following: 

(1) The ANC is directly associated 

with the causative agents, pH and 

dissolved Al, both through empirical 

evidence and mechanistic relationships; 

(2) Empirical evidence shows very 

clear and strong relationships between 

adverse effects and ANC; 

(3) The ANC is a more reliable 

indicator from a modeling perspective, 

allowing use of a body of studies and 

technical analyses related to ANC and 

acidification to inform the development 

of the standard; and 

(4) The ANC literally embodies the 

concept of acidification as posed by the 

basic principles of acid base chemistry 

and the measurement method used to 

estimate ANC and, therefore, serves as 

a direct index to protect against 

acidification. 

Ecological indicators of acidification 

in aquatic ecosystems can be chemical 

or biological components of the 

ecosystem that are altered by the 

acidifying effects of nitrogen and sulfur 

deposition. A desirable ecological 

indicator for aquatic acidification is one 

that is measurable or estimable, linked 

causally to deposition of nitrogen and 

sulfur, and linked causally, either 

directly or indirectly to ecological 

effects known or anticipated to 

adversely affect public welfare. 

As summarized in chapter 2 of the 

PA, atmospheric deposition of NOy and 

SOX causes aquatic acidification 

through the input of strong acid anions 

(e.g., NO3

¥ and SO4

2¥) that ultimately 

shifts the water chemistry equilibrium 

toward increased hydrogen ion levels 

(or decreased pH). The anions are 

deposited either directly to the aquatic 

ecosystem or indirectly via 

transformation through soil nitrification 

processes and subsequent drainage from 

terrestrial ecosystems. In other words, 

when these anions are mobilized in the 

terrestrial soil, they can leach into 

adjacent water bodies. Aquatic 

acidification is indicated by changes in 

the surface water chemistry of 

ecosystems. In turn, the alteration of 

surface water chemistry has been linked 

to negative effects on the biotic integrity 

of freshwater ecosystems. There is a 

suite of chemical indicators that could 

be used to assess the effects of 

acidifying deposition on lake or stream 

acid-base chemistry. These indicators 

include ANC; alkalinity (ALK); base 

neutralizing capacity, commonly 

referred to as acidity (ACY); surface 

water pH; concentrations of trivalent 

aluminum, Al+3; and concentrations of 

major anions (SO4

2¥, NO3

¥), cations 

(Ca2+, Mg+2, K+), or sums of cations or 

anions. 

The ANC and ALK are very similar 

quantities and are used interchangeably 

in the literature and for some of the 

analyses presented in this document. 

Both ANC and ALK are defined as the 

amount of strong acid required to reach 

a specified equivalence point. For acidbase solutions, an equivalence point can 

be thought of as the point at which the 

addition of strong acids (i.e., titration) is 

no longer neutralized by the solution. 

This explains the term acid neutralizing 

capacity, or ANC, as ANC relates 

directly to the capacity of a system to 

neutralize acids. The differences 

between ANC and ALK are based on 

operational definitions and subject to 

various interpretations. The ANC is 

preferred over ALK as the body of 

scientific evidence has focused on ANC 

and effects relationships. The ALK is 

more widely associated with more 

general characterizations of water 

quality such as the relative hardness of 

water associated with carbonates. 

Indictors such as the concentrations 

of specific anions, cations, or their 

groupings, while relevant to 

acidification processes, are not robust 

acidification indicators as it is the 

relative balance of cations and anions 

that is more directly associated with 

acidification. That balance is captured 

by ANC and ALK. Acidity, ACY, is the 

converse of ANC and indicates how 

much strong base it takes to reach an 

equivalence point. Because ACY is not 

used in most ecosystem assessments, 

the body of information relating ACY to 

effects is too limited to serve as a basis 

for an appropriate ecological indicator. 

Aluminum and other metals are 

causative toxic agents that directly 

impair biological functions. However, 

Al, or metals in general, have high 

variability in concentrations that can be 

linked to effects, often at extremely low 

levels which in some cases approach 

detectability limits, exhibit rapid 

transient responses, and are often 

confounded by the presence of other 

toxic metals. These concerns limit the 

use of metals as reliable and measurable 

ecological indicators. Hydrogen ion (H+) 

concentrations, using their negative 

logarithmic values, or pH, are well 

correlated with adverse effects, as 

discussed above in section II.A, and 

determine the solubility of metals such 

as aluminum. However, pH is not a 

preferred acidification indicator due to 

its highly transient nature and other 

concerns, as discussed below. 

Having reasoned that ANC is a 

preferred indicator to ALK, ACY, 

individual metals or groupings of ions, 

the PA considers the relative merits of 

ANC compared to pH, which is a well 

recognized indicator of acidity and a 

more direct causative agent with regard 

to adverse effects. First, the linkage 

between ANC and pH is considered in 

recognition of the causative association 

between pH and effects. 

The ANC is not the direct causative 

toxic agent impacting aquatic species 

diversity. The scientific literature 

generally emphasizes the links between 

pH and adverse effects as described 

above in section II.A. It is important, 

therefore, to consider the extent to 

which ANC and pH are well related 

from a mechanistic perspective as well 

as through empirical evidence. The 

ANC and pH are co-dependent on each 

other based on the requirement that all 

solutions are electrically neutral, 

meaning that any solution must satisfy 

the condition that all negatively charged 

species must be balanced by all 

positively charged species. The ANC is 

defined as the difference between strong 

anions and cations (US EPA, 2011, 

equation 7–13). 

While the chemistry can be complex, 

the co-dependency between ANC and 

pH is explained by recognizing that 

positively charged hydrogen, H+, is 

incorporated in the charge balance 

relationships related to the overall 

solution chemistry which also defines 

ANC. The positive, directional codependency (i.e., ANC and pH increase 

together) is further explained in concept 

as ANC reflects how much strong acid 

(i.e., how much hydrogen ion) it takes 

to titrate to an equivalence point. Strong 

observed correlations between pH and 

ANC as described in the PA support 

these mechanistic relationships. 

As discussed above in section II.A, 

there are well established examples of 

ANC correlating strongly with a variety 

of ecological effects which are 

summarized in the PA (US EPA, 2011, 

Table 3–1). Because pH and ANC are 

well correlated and linearly dependent 

over the pH ranges (4.5–6) where 

adverse ecological effects are observed, 

evidence of clear associations exist 

between ANC and adverse ecological 

effects as described in the PA. In large 

measure, this dependence between pH 


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and ANC and the relationship of both 

pH and ANC to effects, speak directly to 

the appropriateness of ANC with respect 

to its use as an ecological indicator. 

Thus, there is a clear association 

between ANC and ecological effects, 

although there is a more direct causal 

relationship between pH and ecological 

effects. Nonetheless, ANC is preferred as 

an ecological indicator based on its 

superior ability to provide a linkage 

with deposition in a meaningful and 

quantifiable manner, a role that is 

served far more effectively by ANC than 

by pH. While both ANC and pH are 

clearly associated with the effects of 

concern, ANC is superior in linking 

these effects to deposition. 

The PA notes that the basis for this 

conclusion is that acidifying 

atmospheric deposition of nitrogen and 

sulfur is a direct input of potential 

acidity (ACY), or, in terms of ANC, such 

deposition is relevant to the major 

anions that reduce the capacity of a 

water body to neutralize acidity. 

Consequently, there is a well defined 

linear relationship between potential 

acidifying deposition and ANC. This 

ANC-deposition relationship facilitates 

the linkage between ecosystem models 

that calculate an ecological indicator 

and the atmospheric deposition of NOy

and SOX. On the other hand, there is no 

direct linear relationship between 

deposition and pH. While acid inputs 

from deposition lower pH, the 

relationship can be extremely nonlinear 

and there is no direct connection from 

a modeling or mass balance perspective 

between the amount of deposition 

entering a system and pH. The term 

‘‘mass balance’’ underlies the basic 

formulation of any physical modeling 

construct, for atmospheric or aquatic 

systems, and refers to the accounting of 

the flow of mass into a system, the 

transformation to other forms, and the 

loss due to flow out of a system and 

other removal processes. The ANC is a 

conserved property. This means that 

ANC in a water body can be accounted 

for by knowledge of how much ANC 

initially exists, how much flows in and 

is deposited, and how much flows out. 

In contrast, hydrogen ion concentration 

in the water, the basis for pH, is not a 

conserved property as its concentration 

is affected by several factors such as 

temperature, atmospheric pressure, 

mixing conditions of a water body, and 

the levels of several other chemical 

species in the system. The disadvantage 

of pH lacking conservative properties is 

that there is a very complex connection 

between changes in ambient air 

concentrations of NOy and SOX and pH. 

The discussion of basic water 

chemistry of natural systems in chapter 

2 of the PA provides further details on 

why pH is not a conserved quantity and 

is subject to rapid transient response 

behavior that makes it difficult to use as 

a reliable and functional ecological 

indicator. The observed pH-to-ANC 

relationship (US EPA, 2011, figure 7–2) 

partially explains the concern with pH 

responding too abruptly. In the region 

where pH ranges roughly from 4.5 to 6 

and is of greatest relevance to effects 

(US EPA, 2011, figure 7–4), there clearly 

is more sensitivity of pH to changes in 

ANC in the ANC range from 

approximately 0 to 50 μeq/L. A focus on 

this part of the ANC-to-pH relationship 

shows that ANC associates well with pH 

in a fairly linear manner. However, the 

pH range from 4.5 to 6 also includes one 

of the very steepest parts of the slope 

relating pH as a function of ANC, where 

ANC ranges down below 0 μeq/L, which 

is subject to very rapid change in ANC, 

or deposition inputs. This part of the 

relationship coincides with reduced 

levels of ANC and hence with reduced 

ability to neutralize acids and moderate 

pH fluctuations. This response behavior 

can be extended to considering how pH 

would change in response to deposition, 

or ambient concentrations, of NOy and 

SOX, which can be viewed as ‘‘ANClike’’ inputs. 

In summary, because ANC clearly 

links both to biological effects of aquatic 

acidification as well as to acidifying 

inputs of NOy and SOX deposition, the 

PA concludes that ANC is an 

appropriate ecological indicator for 

relating adverse aquatic ecosystem 

effects to acidifying atmospheric 

deposition of SOx and NOy, and is 

preferred to other potential indicators. 

In reaching this conclusion, the PA 

notes that in its review of the first draft 

PA, CASAC concluded that 

‘‘information on levels of ANC 

protective to fish and other aquatic biota 

has been well developed and presents 

probably the lowest level of uncertainty 

in the entire methodology’’ (Russell and 

Samet, 2010a). In its more recent review 

of the second draft PA, CASAC agreed 

‘‘that acid neutralizing capacity is an 

appropriate ecological measure for 

reflecting the effects of aquatic 

acidification’’ (Russell and Samet, 

2010b; p. 4). 

2. Linking ANC to Deposition 

There is evidence to support a 

quantified relationship between 

deposition of nitrogen and sulfur and 

ANC. This relationship was analyzed in 

the REA for two case study areas, the 

Adirondack and Shenandoah 

Mountains, based on time-series 

modeling and observed trends. In the 

REA analysis, long-term trends in 

surface water nitrate, sulfate and ANC 

were modeled using MAGIC for the two 

case study areas. These data were used 

to compare recent surface water 

conditions in 2006 with preindustrial 

conditions (i.e. preacidification 1860). 

The results showed a marked increase 

in the number of acid impacted lakes, 

characterized as a decrease in ANC 

levels, since the onset of anthropogenic 

nitrogen and sulfur deposition, as 

discussed in chapter 2 of the PA. 

In the REA, more recent trends in 

ANC, over the period from 1990 to 2006, 

were assessed using monitoring data 

collected at the two case study areas. In 

both case study areas, nitrate and sulfate 

deposition decreased over this time 

period. In the Adirondack Mountains, 

this corresponded to a decreased 

concentration of nitrate and sulfate in 

the surface waters and an increase in 

ANC (U.S. EPA, 2009, section 4.2.4.2). 

In the Shenandoah Mountains, there 

was a slight decrease in nitrate and 

sulfate concentration in surface waters 

corresponding to modest increase in 

ANC from 50 μeq/L in 1990 to 67 μeq/ 

L in 2006 (U.S. EPA, 2009, section 

4.2.4.3, Appendix 4, and section 3.4). 

In the REA, the quantified 

relationship between deposition and 

ANC was investigated using ecosystem 

acidification models, also referred to as 

acid balance models or critical loads 

models (U.S. EPA, 2011, section 2 and 

U.S. EPA, 2009, section 4 and Appendix 

4). These models quantify the 

relationship between deposition of 

nitrogen and sulfur and the resulting 

ANC in surface waters based on an 

ecosystem’s inherent generation of ANC 

and ability to neutralize nitrogen 

deposition through biological and 

physical processes. A critical load is 

defined as the amount of acidifying 

atmospheric deposition of nitrogen and 

sulfur beyond which a target ANC is not 

reached. Relatively high critical load 

values imply that an ecosystem can 

accommodate greater deposition levels 

than lower critical loads for a specific 

target ANC level. Ecosystem models that 

calculate critical loads form the basis for 

linking deposition to ANC. 

As discussed in chapter 2 of the PA, 

both dynamic and steady state models 

calculate ANC as a function of 

ecosystem attributes and atmospheric 

nitrogen and sulfur deposition, and can 

be used to calculate critical loads. 

Steady state models are time invariant 

and reflect the long term consequences 

associated with an ecosystem reaching 

equilibrium under a constant level of 

atmospheric deposition. Dynamic 

models are time variant and take into 

account the time dependencies inherent 

in ecosystem hydrology, soil and 


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7This section discusses the linkages between 

deposition of nitrogen and sulfur and ANC. Section 

III.B.3 then discusses the linkages between 

atmospheric concentrations of NOY and SOX and 

deposition of nitrogen and sulfur. 8Because Neco is only relevant to nitrogen 

deposition, in rare cases where Neco is greater than 

the total nitrogen deposition, the critical load 

would be defined only in terms of acidifying 

deposition of sulfur and the Neco term in equation 

III–1 would be set to zero. 

biological processes. Dynamic models 

like MAGIC can provide the time series 

response of ANC to deposition whereas 

steady state models provide a single 

ANC relationship to any fixed 

deposition level. Dynamic models 

naturally are more complex than steady 

state models as they attempt to capture 

as much of the fundamental 

biogeochemical processes as practicable, 

whereas steady state models depend on 

far greater parameterization and 

generalization of processes that is 

afforded, somewhat, by not having to 

accounting for temporal variability. 

The PA notes that steady state models 

are capable of addressing the question of 

what does it take to reach and sustain 

a specific level of ANC. Dynamic 

models are also capable of addressing 

that question, but can also address the 

question of how long it takes to achieve 

that result. Dynamic models afford the 

ability for more comprehensive 

treatment of a variety of processes 

throughout the surface, soil and bedrock 

layers within an ecosystem. For 

example, steady state models treat 

sulfate as a mobile anion throughout the 

system, meaning that the sulfate that is 

deposited to a watershed enters the 

water column and is not influenced by 

soil adsorption or cation exchange. 

Dynamic models can incorporate these 

time variant processes. The use of a 

steady state model treating sulfate as 

totally mobile does not necessarily 

conflict with the possibility of sulfate 

acting as a less than mobile ion at 

certain times. The steady state 

assumption is premised on the long 

term behavior of sulfate which can 

undergo periods of net adsorption 

followed by periods of net desorption 

which can balance out over time. The 

PA recognizes that as the richness of the 

available data increases, in terms of 

parameters and spatial resolution, the 

incorporation of dynamic modeling 

approaches in the standard setting 

process should become more feasible. In 

determining an appropriate modeling 

approach for the development of a 

NAAQS in this review, the PA considers 

both the relevance of the question 

addressed as well as the ability to 

perform modeling that provides relevant 

information for geographic areas across 

the country. 

Dynamic models require a large 

amount of catchment level-specific data 

relative to steady state models. Because 

of the time invariant nature of steady 

state models, the data requirements that 

integrate across a broad spectrum of 

ecosystem processes is achievable and 

available now at the national level. 

Water quality data to support steady 

state models currently exist for 

developing a national data base for 

modeling nearly 10,000 catchments in 

the contiguous U.S. In contrast, the data 

needs to support dynamic models for 

national-scale analyses simply are not 

available at this time. Further, the 

information provided by steady state 

modeling would be sufficient to develop 

and analyze alternative NAAQS and the 

kind of protection they would afford. 

While it would be of interest to also 

obtain information about how much 

time it would take for a target ANC level 

to be achieved, the absence of such 

information does not preclude 

developing and evaluating alternative 

NAAQS using the AAI structure. Based 

on the above considerations, the PA 

concludes that at this time steady state 

critical load modeling is an appropriate 

tool for linking long-term ANC levels to 

atmospheric deposition of nitrogen and 

sulfur for development of an AAI that 

has national applicability. 

A steady state model is used to define 

the critical load, which is the amount of 

atmospheric deposition of nitrogen (N) 

and sulfur (S) beyond which a target 

ANC is not achieved and sustained.7 It 

is expressed as: 

Where: 

CLANClim(N + S) is the critical load of 

deposition, with units of equivalent 

charge/(area-time); 

[BC]0

¥* is the natural contribution of base 

cations from weathering, soil processes and 

preindustrial deposition, with units of 

equivalent charge/volume; 

[ANClim] is the target ANC value, with units 

of equivalent charge/volume; Q is the 

catchment level runoff rate governed by 

water mass balance and dominated by 

precipitation, with units of distance/ 

time; and 

Neco is the amount of nitrogen deposition 

that is effectively neutralized by a variety 

of biological (e.g., nutrient uptake) and 

physical processes, with units of 

equivalent charge/(area-time). 

Equation III–1 is a modified 

expression that adopts the basic 

formulation of the steady state models 

that are described in chapter 2 of the 

PA. More detailed discussion of the 

rationale, assumptions and derivation of 

equation III–1, as well as all of the 

equations in this section, are included 

in Appendix B of the PA. The equation 

simply reflects the amount of deposition 

of nitrogen and sulfur from the 

atmosphere, CLANClim(N + S), that is 

associated with a sustainable long-term 

ANC target, [ANClim], given the capacity 

of the natural system to generate ANC, 

[BC]0

¥*, and the capacity of the natural 

system to neutralize nitrogen 

deposition, Neco. This expression of 

critical load is valid when nitrogen 

deposition is greater than Neco.8 The 

runoff rate, Q, allows for balancing mass 

in the two environmental mediums— 

atmosphere and catchment. This critical 

load expression can be focused on a 

single water system or more broadly. To 

extend applicability of the critical load 

expression (equation III–1) from the 

catchment level to broader spatial areas, 

the terms Qr and CLr, are used, which 

are the runoff rate and critical load, 

respectively, of the region over which 

all the atmospheric terms in the 

equation are defined. 

In considering the contributions of 

SOx or NOy species to acidification, it is 

useful to think of every depositing 

nitrogen atom as supplying one 

equivalent charge unit and every sulfur 

atom as depositing two charge units. 

The PA uses equivalent charge per 

volume as a normalizing tool in place of 

the more familiar metrics such as mass 

or moles per volume. This allows for a 

clearer explanation of many of the 

relationships between atmospheric and 

ecosystem processes that incorporate 

mass and volume unit conventions 

somewhat specific to the environmental 

media of concern (e.g., m3 for air and 

liter for liquid water). Equivalent charge 

reflects the chemistry equilibrium 

fundamentals that assume 

electroneutrality, or balancing charge 

where the sum of cations always equals 

the sum of anions. 

As presented above, the terms S and 

N in the CLANClim (N + S) term broadly 

represent all species of sulfur or 

nitrogen that can contribute to 


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acidifying deposition. This follows 

conventions used in the scientific 

literature that addresses critical loads, 

and it reflects all possible acidifying 

contributions from any sulfur or 

nitrogen species. For all practical 

purposes, S reflects SOx as described 

above, the sum of sulfur dioxide gas and 

particulate sulfate. However, N in 

equation III–1 includes both oxidized 

forms, consistent with the ambient 

indicator, NOy, in addition to the 

reduced nitrogen species, ammonia and 

ammonium ion, referred to as NHx. The 

NHX is included in the critical load 

formulation because it contributes to 

potentially acidifying nitrogen 

deposition. Consequently, from a mass 

balance or modeling perspective, the 

form of the standard needs to account 

for NHX, as described below. 

3. Linking Deposition to Ambient Air 

Indicators 

The last major component of the form 

illustrated in Figure III–1 addresses the 

linkage between deposition of nitrogen 

and sulfur and concentrations of the 

ambient air indicators, NOY and SOX. 

To link ambient air concentrations with 

deposition, the PA defines a 

transference ratio, T, as the ratio of total 

wet and dry deposition to ambient 

concentration, consistent with the area 

and time period over which the 

standard is defined. To express 

deposition of NOY and SOX in terms of 

NOY and SOX ambient concentrations, 

two transference ratios were defined, 

where TSOx equals the ratio of the 

combined dry and wet deposition of 

SOx to the ambient air concentration of 

SOx, and TNOY equals the ratio of the 

combined dry and wet deposition of 

NOY to the ambient air concentration of 

NOY. 

As described in chapter 7 of the PA, 

reduced forms of nitrogen (NHx) are 

included in total nitrogen in the critical 

load equation, III–1. Reduced forms of 

nitrogen are treated separately, as are 

NOy and SOx, and the transference 

ratios are applied. This results in the 

following critical load expression that is 

defined explicitly in terms of the 

indicators NOY and SOx: 

This is the same equation as III–1, with 

the deposition associated with the 

critical load translated to deposition 

from ambient air concentrations via 

transference ratios. In addition, 

deposition of reduced nitrogen, 

oxidized nitrogen and oxidized sulfur 

are treated separately. 

Transference ratios are a modeled 

construct, and therefore cannot be 

compared directly to measurable 

quantities. There is an analogy to 

deposition velocity, as a transference 

ratio is basically an aggregated weighted 

average of the deposition velocities of 

all contributing species across dry and 

wet deposition, and transference ratio 

units are expressed as distance/time. 

However, wet deposition commonly is 

not interpreted as the product of a 

concentration times a velocity. Direct 

wet deposition observations are 

available which integrate all of the 

processes, regardless of how well they 

may be understood, related to wet 

deposition into a measurable quantity. 

There are reasonable analogies between 

the processes governing dry and wet 

deposition, from a fundamental mass 

transfer perspective. In both cases there 

is a transfer of mass between the dry 

ambient phase and another medium, 

either a surface or vegetation in the case 

of dry deposition, or a rain droplet or 

cloud in the case of wet precipitation. 

The specific thermodynamic properties 

and chemical/biological reactions that 

govern the transfer of dry mass to plants 

or aqueous droplets differ, but either 

process can be based on conceptualizing 

the product of a concentration, or 

concentration difference, times a mass 

transfer coefficient which is analogous 

to the basic dry deposition model: dry 

deposition = concentration × velocity 

(U.S. EPA, 2011, Appendix F). 

Transference ratios require estimates 

of wet deposition of NOy and SOX, dry 

deposition of NOY and SOX, and 

ambient air concentrations of NOY and 

SOx. Possible sources of information 

include model estimates or a 

combination of model estimates and 

observations, recognizing that dry 

deposition is a modeled quantity that 

can use observed or modeled estimates 

of concentration. The limited amount of 

NOY measurements in acid-sensitive 

areas as well as the combination of 

representative NOY, SO2 and SO4

observations generally preclude the use 

of observations for development of a 

standard that is applicable nationally. 

The PA considers a blending of 

observations and models to take 

advantage of their relative strengths; 

e.g., combining the NADP wet 

deposition observations, modeled dry 

deposition, and a mix of modeled and 

observed concentrations, using the 

model for those species not measured or 

measured with very sparse spatial 

coverage. A potential disadvantage of 

mixing and matching observations and 

model estimates is to lose consistency 

afforded by using just modeling alone. 

A modeling platform like CMAQ is 

based on adhering to consistent 

treatment of mass conservation, by 

linking emission inputs with air 

concentrations and concentrations to 

deposition. Inconsistencies from 

combining processes from different 

analytical platforms increase the chance 

that mass (of nitrogen or sulfur) would 

unintentionally be increased or 

decreased as the internal checking that 

assures mass conservation is lost. 

Transference ratios incorporate a broad 

suite of atmospheric processes and 

consequently an analytical approach 

that instills consistency in the linkage of 

these processes is preferable to an 

approach lacking such inherent 

consistency. This contention does not 

mean that observations alone, if 

available, could not be used, but 

suggests that the inconsistencies in 

combining models and observations for 

the purposes of developing transference 

ratios has the potential for creating 

unintended artifacts. 

While there is a reasonable 

conceptual basis for the concept of an 

aggregated deposition velocity referred 

to in the PA as a transference ratio, there 

is very limited ability to compare 

observed and calculated ratios. This is 

because the deposition velocity is 

dependent on individual species, and 

the mass transfer processes of wet and 

dry removal, while conceptually 

similar, are different. Consequently, 

there does not exist a meaningful 

approach to measure such an aggregated 

or lumped parameter. Therefore, at this 

time, the evaluation of transference 

ratios is based on sensitivity studies, 

analysis of variability, and comparisons 

with other models, as described in 

Appendix F of the PA. 

As discussed in Appendix F, the 

interannual variability, as well as the 

sensitivity to emission changes of 

roughly 50 percent, results in changes of 

transference ratios of approximately 5 to 

10 percent. Part of the reason for this 

inherent stability is due to the codependence of concentration and 

deposition. For example, as 

concentrations are reduced as a result of 

emissions reductions, deposition in turn 


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9Because NHx is characterized directly as 

deposition, not as an ambient concentration in this 

equation, no transference ratio is needed for this 

term. 

is reduced since deposition is a direct 

linear function of concentration leading 

to negligible impact on the depositionto-concentration ratio. Likewise, an 

overestimate of concentration likely 

does not induce a bias in the 

transference ratio. While it is important 

to continue to improve the model’s 

ability to match ambient concentrations 

in time and space, the bias of a modeled 

estimate of concentration relative to 

observations does not necessarily result 

in a bias in a calculated transference 

ratio. In effect, this consideration of bias 

cancellation reduces the sensitivity of 

transference ratios to model 

uncertainties and affords increased 

confidence in the stability of these 

ratios. Based on the series of sensitivity 

and variability analyses, the PA 

concludes that the transference ratios 

are relatively stable and provide a sound 

metric for linking deposition and 

concentration. 

As discussed in the PA, transference 

ratios are dependent on the platform 

upon which they are constructed. 

Comparisons of transference ratios 

constructed from different modeling 

platforms do exhibit significant 

differences. While this divergence of 

results may be explained by a variety of 

differences in process treatments, input 

fields and incommensurabilities in 

species definitions and spatial 

configurations, it does suggest two very 

important conclusions. First, the idea of 

using multiple platforms for different 

parts of the country may be problematic 

as there does not exist a reliable 

approach to judge acceptance which is 

almost always based on comparisons to 

observations. Second, since transference 

ratios are based on concentrations and 

deposition, as the uncertainties in each 

of those components are reduced, the 

relative uncertainty in the ratios also is 

reduced. This means that basic 

improvements in the model’s ability to 

reproduce observed wet deposition and 

ambient concentration fields enhance 

the relative confidence in the 

constructed transference ratios. 

Similarly, as in-situ dry deposition flux 

measurements become available that 

enable a more rigorous evaluation and 

diagnosis of modeled dry deposition 

processes, the expected improved 

treatment of dry deposition also would 

increase confidence in transference 

ratios. Finally, deposition is directly 

related to ambient air concentrations. 

Models like CMAQ rely on the 

concentration-to-deposition linkage to 

calculate deposition, which is the 

foundation for broadly based and robust 

assessments addressing atmospheric 

deposition. In principle, the use of a 

modeled constructed transference ratio 

is based on the same premise by which 

we use models to estimate deposition in 

the first place. 

The shortage of widely available 

ambient air observations and the fact 

that estimates of dry deposition requires 

modeling, collectively suggests that a 

unified modeling platform is the best 

approach for constructing transference 

ratios. The PA (U.S. EPA, 2011, section 

2) considers CMAQ and other models, 

such as CAMx and Canada’s 

AURAMS—A Unified Regional Airquality Modeling System (Smythe et al., 

2008), and concludes that CMAQ is the 

preferred modeling platform for 

constructing transference ratios. This 

conclusion reflects the view that for the 

purposes of defining transference ratios, 

a modeling platform should: (1) Be a 

multiple pollutant model recognizing 

the myriad of connections across 

pollutant categories that directly and 

indirectly impact nitrogen and sulfur 

characterization, (2) include the most 

comprehensive scientific treatments of 

atmospheric processes that relate 

directly and indirectly to characterizing 

concentrations and deposition, (3) have 

an infrastructure capability that 

accommodates the inclusion of 

improved scientific treatments of 

relevant processes and important input 

fields, and (4) undergo frequent reviews 

regarding the adequacy of the 

underlying science as well as the 

appropriateness in applications. The 

CMAQ platform exhibits all these 

characteristics. It has been (and 

continues to be) extensively evaluated 

for several pollutant categories, and is 

supported by a central infrastructure of 

EPA scientists, whose mission is to 

improve and evaluate the CMAQ 

platform. More directly, CMAQ, and its 

predecessor versions, has a long track 

record going back to the NAPAP in the 

1980s of specific improvements in 

deposition processes, which are 

described in Appendix F of the PA. 

4. Aquatic Acidification Index 

Having established the various 

expressions that link atmospheric 

deposition of nitrogen and sulfur to 

ANC and the transference ratios that 

translate atmospheric concentrations to 

deposition of nitrogen and sulfur, the 

PA derived the following expression of 

these linkages, which separates reduced 

forms of nitrogen, NHX, from oxidized 

forms: 

Equation III–3 is the basic expression 

of the form of a standard that translates 

the conceptual framework into an 

explicit expression that defines ANC as 

a function of the ambient air indicators, 

NOY and SOX reduced nitrogen 

deposition,9 and the critical load 

necessary to achieve a target ANC level. 

This equation calculates an expected 

ANC value based on ambient 

concentrations of NOY and SOX. The 

calculated ANC will differ from the 

target ANC (ANClim) depending on how 

much the nitrogen and sulfur deposition 

associated with NOY, SOX, and NHX

differs from the critical load associated 

with just achieving the target ANC. 

Based on equation III–3, the PA 

defines an AAI that is more simply 

stated using terms that highlight the 

ambient air indicators: 

where the AAI represents the long term 

(or steady state) ANC level associated 

with ambient air concentrations of NOY

and SOX. The factors F1 through F4 

convey three attributes: a relative 

measure of the ecosystem’s ability to 

neutralize acids (F1), the acidifying 

potential of reduced nitrogen deposition 

(F2), and the deposition-toconcentration translators for NOY (F3) 

and SOX (F4). 

Specifically: 

F1 = ANClim + CLr/Qr; 


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10We note that an 85th area within Omernik’s 

Ecoregion Level III is currently being developed for 

California. 

F2 = NHx/Qr = NHx deposition divided by 

Qr; 

F3 = TNOy/ Qr; TNOy is the transference ratio 

that converts ambient air concentrations 

of NOy to deposition of NOy; and 

F4 = TSOx/ Qr; TSOx is the transference ratio 

that converts ambient air concentrations 

of SOX to deposition of SOX. 

All of these factors include 

representative Qr to maintain unit (and 

mass) consistency between the AAI and 

the terms on the right side of equation 

III–4. 

The F1 factor is the target ANC level 

plus the amount of deposition (critical 

load) the ecosystem can receive and still 

achieve the target level. It incorporates 

an ecosystem’s ability to generate acid 

neutralizing capacity through base 

cation supply ([BC]*0) and to neutralize 

acidifying nitrogen deposition through 

Neco, both of which are incorporated in 

the CL term. As noted above, because 

Neco can only neutralize nitrogen 

deposition (oxidized or reduced) there 

may be rare cases where Neco exceeds 

the combination of reduced and 

oxidized nitrogen deposition. 

Consequently, to ensure that the AAI 

equation is applicable in all cases that 

may occur, equation III–4 is conditional 

on total nitrogen deposition, {NHX + 

F3[NOy]}, being greater than Neco. In 

rare cases where Neco is greater than 

{NHX + F3[NOy]}, F2, F3, and Neco 

would be set equal to 0 in the AAI 

equation. The consequence of setting F2 

and F3 to zero is simply to constrain the 

AAI calculation just to SOx, as nitrogen 

would have no bearing on acidifying 

contributions in this case. 

The PA concludes that equation III–4 

(U.S. EPA, 2011,equation 7–12), which 

defines the AAI, is ecologically relevant 

and appropriate for use as the form of 

a national standard designed to provide 

protection for aquatic ecosystems from 

the effects of acidifying deposition 

associated with concentrations of oxides 

of nitrogen and sulfur in the ambient 

air. This AAI equation does not, 

however, in itself, define the spatial 

areas over which the terms of the 

equation would apply. To specify values 

for factors F1 through F4, it is necessary 

to define spatial areas over which these 

factors are determined. Thus, it is 

necessary to identify an approach for 

spatially aggregating water bodies into 

ecologically meaningful regions across 

the U.S., as discussed below. 

5. Spatial Aggregation 

As discussed in the PA, one of the 

unique aspects of this form is the need 

to consider the spatial areas over which 

values for the F factors in the AAI 

equation are quantified. Ecosystems 

across the U.S. exhibit a wide range of 

geological, hydrological and vegetation 

characteristics that influence greatly the 

ecosystem parameters, Q, BC0

¥* and 

Neco that are incorporated in the AAI. 

Variations in ecosystem attributes 

naturally lead to wide variability in the 

sensitivities of water bodies in the U.S. 

to acidification, as well as in the 

responsiveness of water bodies to 

changes in acidifying deposition. 

Consequently, variations in ecosystem 

sensitivity, and the uncertainties 

inherent in characterizing these 

variations, must be taken into account 

in developing a national standard. In 

developing a secondary NAAQS to 

protect public welfare, the focus of the 

PA is on protecting sensitive 

populations of water bodies, not on each 

individual water body, which is 

consistent with the Agency’s approach 

to protecting public health through 

primary NAAQS that focus on 

susceptible populations, not on each 

individual. 

The approach used for defining 

ecologically relevant regions across the 

U.S. in the PA (U.S. EPA, 2011, section 

7.2.5) is described below, along with 

approaches to characterizing each 

region as acid sensitive or relatively 

non-acid sensitive. This characterization 

facilitates a more detailed analysis and 

focus on those regions that are relatively 

more acid sensitive. This 

characterization is also used to avoid 

over-protection in relatively non-acid 

sensitive regions, regions that would 

receive limited benefit from reductions 

in the deposition of oxides of nitrogen 

and sulfur with respect to aquatic 

acidification effects. Approaches to 

developing representative values for 

each of the terms in the AAI equation 

(factors F1 through F4) for each 

ecologically relevant region for which 

sufficient data are available are then 

discussed. These spatial aggregation 

approaches are generally applicable to 

the contiguous U.S. The following 

discussion also addresses the 

development of factors for data-limited 

regions and specifically for Hawaii, 

Alaska and the U.S. territories. 

Stated more simply, this section 

discusses appropriate ways to divide the 

country into ecologically relevant 

regions; to characterize each region as 

either acid sensitive or relatively nonacid sensitive; and to determine values 

of factors F1 through F4 for each region, 

taking into consideration the acid 

sensitivity of each region and the 

availability of relevant data. For each 

such region, the AAI would be 

calculated based on the values of factors 

F1 through F4 specified for that region. 

In considering approaches to spatial 

aggregation, the PA focuses on methods 

that have been developed to define 

ecologically relevant regions, referred to 

as ecoregions, which are meaningfully 

related to the factors that are relevant to 

aquatic acidification. As noted above, 

the PA did not focus on looking at each 

individual water body, nor did it focus 

on aggregating over the entire nation, 

which would preclude taking into 

account the inherent variability in 

atmospheric and ecological factors that 

fundamentally modify the relationships 

that are central to the development of an 

ecologically relevant AAI. 

Based on considering available 

classification schemes, the PA 

concludes that Omernik’s ecoregion 

classification (as described at http:// 

www.epa.gov/wed/pages/ecoregions) is 

the most appropriate method to 

consider for the purposes of this review. 

This classification offers several levels 

of spatial delineation, has undergone an 

extensive scientific peer review process, 

and has explicitly been applied to 

delineating acid sensitive areas within 

the U.S. Further, the PA concludes that 

ecoregion level III (Figure III–1) 

resolution, with 84 defined ecoregions 

in the contiguous U.S.,10 is the most 

appropriate level to consider for this 

purpose. The spatial resolution afforded 

by level III strikes an appropriate 

balance relative to the reasoning that 

supports conclusions on indicators, as 

discussed above. The PA concludes that 

the most detailed level of resolution 

(level IV) is not appropriate given the 

limited data availability to address 

nearly 1,000 subdivisons within that 

level and the currently evolving nature 

of level IV regions. Further, level III 

ecoregions are preferred to level II in 

that level III ecoregions, but not level II 

ecoregions, are largely contiguous in 

space which allows for a more coherent 

development of information to quantify 

the AAI factors and to characterize the 

concentrations of NOy and SOx in the 

ambient air within each ecoregion. 

Appendix C of the PA includes a 

description of each level III ecoregion. 

The PA notes that the use of ecoregions 

is an appropriate spatial aggregation 

scheme for an AAI-based standard 

focused on deposition-related aquatic 

acidification effects, while many of the 

same ecoregion attributes may be 

applicable in subsequent NAAQS 

reviews that may address other 

deposition-related aquatic and 

terrestrial ecological effects. Because 

atmospheric deposition is modified by 

ecosystem attributes, the types of 

vegetation, soils, bedrock geology, and 


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topographic features that are the basis of 

this ecoregion classification approach 

also will likely be key attributes for 

other deposition-related effects (e.g., 

terrestrial acidification, nutrient 

enrichment) that link atmospheric 

concentrations to an aquatic or 

terrestrial ecological indicator. 

a. Ecoregion Sensitivity 

The PA used Omernik’s original 

alkalinity data (U.S. EPA, 2011, section 

2) and more recent ANC data to 

delineate two broad groupings of 

ecoregions: Acid-sensitive and relatively 

non-acid sensitive ecoregions. This 

delineation was made to facilitate 

greater focus on those ecoregions with 

water bodies that generally have greater 

acid sensitivity and to avoid overprotection in regions with generally less 

sensitive water bodies. The approach 

used to delineate acid-sensitive and 

relatively non-acid sensitive regions 

included an initial numerical-based 

sorting scheme using ANC data, which 

categorized ecoregions with relatively 

high ANC values as being relatively 

non-acid sensitive. This initial 

delineation resulted in 29 of the 84 

Omernik ecoregions being categorized 

as acid sensitive. Subsequently, land 

use data were also considered to 

determine to what extent an ecoregion 

is of a relatively pristine and rural 

nature by quantifying the degree to 

which active management practices 

related to development and agriculture 

occur in each ecoregion. 

The overall objective is to produce a 

logical and practical grouping of 

ecoregions that experience adverse 

conditions with respect to aquatic 

acidification and are likely to respond to 

changes in concentrations of NOy and 

SOx in the ambient air and to the related 

deposition levels. To achieve this goal, 

a two-step process has been applied, 

first identifying acid sensitive 

ecoregions based on water quality data 

alone, and second identifying among 

those acid-sensitive ecoregions those 

with highly developed and managed 

areas. These highly developed and 

managed ecoregions are placed in a nonacid sensitive category to avoid over 

protection beyond what is requisite to 

protect public welfare. More 

specifically, in determining an 

ecoregion’s acid sensitivity status in 

step 1, ANC data across the 84 

ecoregions are sorted (U.S. EPA, 2011, 

section 7) to determine the number of 

water bodies within a region with ANC 

values suggestive of acid sensitivity, so 

as to screen out regions with an 

overabundance of high ANC values. In 

reviewing the ANC data, the PA 

identified 29 ecoregions that meet two 

criteria: (1) Greater than 5 percent of 

water bodies with data with ANC values 

less than 200 μeq/L and (2) greater than 

1 percent of water bodies with ANC 

values less than 100 μeq/L. In step 2, 

land use data were used to identify 

those acid sensitive ecoregions with 

significant managed areas that would 

not be considered as having a relatively 

pristine and rural character. The 

percentage of the combination of 

developed (residential, transportation, 

industrial and commercial) and 

agricultural (croplands, pastures, 

orchards, vineyards) land use was used 

as an indicator of managed land use 

area. Forest cover was used as an 

indicator of non-managed land use more 

directly reflecting the pristine quality of 

a region. Based on the 2006 National 

Land Cover Data base (NLCD, http:// 

www.epa.gov/mrlc/nlcd-2006.html), 

acid sensitive ecoregions would meet 

both of the following land use data 


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criteria: Percent of developed and 

agricultural area less than 20 percent 

combined with forested area greater 

than 50 percent. The combination of 

steps 1 and 2 identify 22 relatively acid 

sensitive areas (Table III–1 and Figure 

III–2). 

Consideration was also given to the 

use of naturally acidic conditions in 

defining relatively non-acid sensitive 

areas. For example, several of the 

ecoregions located in plains near the 

coast exhibit elevated dissolved organic 

carbon (DOC) levels, which is associated 

with naturally acidic conditions. The 

DOC in surface waters is derived from 

a variety of weak organic acid 

compounds generated from the natural 

availability and decomposition of 

organic matter from biota. 

Consequently, high DOC is associated 

with ‘‘natural’’ acidity, with the 

implication that a standard intended to 

protect against atmospheric 

contributions to acidity is not an area of 

focus. The evidence suggests that 

several of the more highly managed 

ecoregions in coastal or near coastal 

transition zones are associated with 

relatively high DOC values, typically 

exceeding on average 5 mg/l, compared 

to other acid sensitive areas. Although 

there is sound logic to interpret 

naturally acidic areas as relatively nonacid sensitive, natural acidity indicators 

were not explicitly included in defining 

relatively non-acid sensitive areas as 

there does not exist a consensus-based 

quantifiable scientific definition of 

natural acidity. Approaches to explicitly 

define natural acidity likely will be 

pursued in future reviews of the 

standard. 

TABLE III–1—LIST OF 22 ACIDSENSITIVE AREAS

Ecoregion name Ecoregion 

No. 

Ridge and Valley ........................ 8.4.1 

Northern Appalachian Plateau 

and Uplands ............................ 8.1.3 

Piedmont ..................................... 8.3.4 

TABLE III–1—LIST OF 22 ACIDSENSITIVE AREAS—Continued 

Ecoregion name Ecoregion 

No. 

Western Allegheny Plateau ........ 8.4.3 

Southwestern Appalachians ....... 8.4.9 

Boston Mountains ....................... 8.4.6 

Blue Ridge .................................. 8.4.4 

Ouachita Mountains .................... 8.4.8 

Central Appalachians ................. 8.4.2 

Northern Lakes and Forests ....... 5.2.1 

Maine/New Brunswick Plains 

and Hills .................................. 8.1.8 

North Central Appalachians ....... 5.3.3 

Northern Appalachian and Atlantic Maritime Highlands ............ 5.3.1 

Columbia Mountains/Northern 

Rockies ................................... 6.2.3 

Middle Rockies ........................... 6.2.10 

Wasatch and Uinta Mountains ... 6.2.13 

North Cascades .......................... 6.2.5 

Cascades .................................... 6.2.7 

Southern Rockies ....................... 6.2.14 

Sierra Nevada ............................. 6.2.12 

Idaho Batholith ............................ 6.2.15 

Canadian Rockies ...................... 6.2.4 


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11The distribution of critical loads was based on 

CL values calculated with Neco at the lake level. 

Consideration could also be given to using a 

distribution of CLs without Neco and adding the 

ecoregion average Neco value to the nth percentile 

critical load. This would avoid cases where the 

lake-level Neco value potentially could be greater 

than total nitrogen deposition. The CL at the lake 

level represents the CL for the lake to achieve the 

specified national target ANC value. 

12The PA judged the data to be sufficient for this 

purpose if data are available from more than 10 

water bodies in an ecoregion. 

b. Representative Ecoregion-Specific 

Factors 

Having concluded that the Omernik 

level III ecoregions are an appropriate 

approach to spatial aggregation for the 

purpose of a standard to address 

deposition-related aquatic acidification 

effects, the PA uses those ecoregions to 

define each of the factors in the AAI 

equation. As discussed below, factors F1 

through F4 in equation III–4 are defined 

for each ecoregion by specifying 

ecoregion-specific values for each factor 

based on monitored or modeled data 

that are representative of each 

ecoregion. 

i. Factor F1 

As discussed above, factor F1 reflects 

a relative measure of an ecosystem’s 

ability to neutralize acidifying 

deposition, and is defined as: F1 = 

ANClim + CLr/Qr. The value of F1 for 

each ecoregion would be based on a 

representative critical load for the 

ecoregion (CLr) associated with a single 

national target ANC level (ANClim, 

discussed below in section III.D), as 

well as on a representative runoff rate 

(Qr). To specify ecoregion-specific 

values for the term Qr, the PA used the 

median value of the distribution of Q 

values that are available for water 

bodies within each ecoregion. To 

specify ecoregion-specific representative 

values for the term CLr in factor F1, a 

distribution 11 of calculated critical 

loads was created for the water bodies 

in each ecoregion for which sufficient 

water quality and hydrology data are 

available.12 The representative critical 

load was then defined to be a specific 

percentile of the distribution of critical 

loads in the ecoregion. Thus, for 

example, using the 90th percentile 

means that within an ecoregion, 90 

percent of the water bodies would be 

expected to have higher calculated 

critical loads than the representative 

critical load. That is, if the 

representative critical load were to 

occur across the ecoregion, 90 percent of 

the water bodies would be expected to 

achieve the national ANC target or 

better. 

The specific percentile selected as 

part of the definition of F1 is an 

important parameter that directly 

impacts the representative critical load 

specified for each ecoregion, and 

therefore the degree of protectiveness of 

the standard. A higher percentile 

corresponds to a lower critical load and, 

therefore, to lower allowable ambient air 

concentrations of NOy and SOx and 

related deposition to achieve a target 

AAI level. In conjunction with the other 

terms in the AAI equation, alternative 

forms can be appropriately 

characterized in part by identifying a 

range of alternative percentiles. The 

choice of an appropriate range of 

percentiles to consider for acid-sensitive 

and relatively non-acid sensitive 

ecoregions, respectively, is discussed 

below. 

a. Acid-Sensitive Ecoregions 

In identifying percentiles that are 

appropriate to consider for the purpose 

of calculating factor F1 for ecoregions 

characterized as acid sensitive, the PA 

concludes that it is appropriate to focus 

on the lower (more sensitive) part of the 

distribution of critical loads, so as to 

ensure that the ecoregion would be 

represented by relatively more acid 

sensitive water bodies within the 

ecoregion. Specifying factor F1 in this 

way would help to define a standard 

that would be protective of the 

population of acid sensitive water 

bodies within an ecoregion, recognizing 

that even ecoregions characterized as 

acid sensitive may contain a number of 

individual water bodies that are not acid 

sensitive. The PA recognizes that there 

is no basis for independently evaluating 

the degree of protectiveness afforded by 

any specific percentile value, since it is 

the combination of form and level, in 

conjunction with the indicator and 

averaging time, which determine the 

degree of protectiveness of a standard. 

In light of this, the PA concludes that it 

is appropriate to consider a range of 

percentiles, from well above the 50th 

percentile, or median, of the 

distribution to somewhat below the 

highest value (in terms of sensitivity; a 

high degree of sensitivity corresponds to 

a low value for critical load). More 

specifically, the PA concludes it is 

appropriate to consider percentiles in 

the range of the 70th to the 90th 

percentile (of sensitivity). This 

conclusion is based on the judgment 

that it would not be appropriate to 

represent an ecoregion with the lowest 

or near lowest critical load, so as to 

avoid potential extreme outliers that can 

be seen to exist at the extreme end of the 

data distributions, which would not be 

representative of the population of acid 

sensitive water bodies within the 

ecoregion and could lead to an overly 

protective standard. At the same time, 

in considering ecoregions that are 

inherently acid sensitive, it is judged to 

be appropriate to limit the lower end of 

the range for consideration to the 70th 

percentile, a value well above the 

median of the distribution, so that a 

substantial majority of acid-sensitive 

water bodies are protected. 

In considering this conclusion, the 

CASAC Panel noted that the data bases 

for calculating critical loads within an 

ecoregion are not necessarily 

representative of all water bodies within 

an ecoregion. That is, in many 

ecoregions the lake sampling design 

used in studies that generated the 

relevant data may have focused on the 

relatively more sensitive water bodies 

within an ecoregion (Russell and Samet, 

2011a). Consequently, a given percentile 

of the distribution of calculated critical 

loads, based on sampled water bodies, 

may not be representative of that 

percentile of all water bodies across an 

entire ecoregion. To the extent that the 

sampling of water bodies within an 

ecoregion was skewed toward the 

relatively more sensitive water bodies, 

selecting a given percentile from the 

distribution of available critical loads 

would result in a somewhat higher 

percentile of all water bodies within 

that ecoregion having a higher 

calculated critical load than the 

representative critical load value. Thus, 

the extent to which study sampling 

designs have resulted in skewed 

distributions of calculated critical loads 

is an uncertainty that is appropriate to 

consider in selecting a percentile for the 

purpose of defining the factor F1 in the 

AAI equation. 

b. Non-Acid Sensitive Ecoregions 

With regard to identifying percentiles 

that are appropriate to consider for the 

purpose of calculating factor F1 for 

ecoregions characterized as relatively 

non-acid sensitive, the PA recognizes 

that while such ecoregions are generally 

less sensitive to acidifying deposition 

from oxides of nitrogen and sulfur, they 

may contain a number of water bodies 

that are acid sensitive. This category 

includes ecoregions that are well 

protected from acidification effects due 

to natural production of base cations 

and high ANC levels, as well as 

naturally acidic systems with limited 

base cation production and 

consequently very low critical loads. 

Therefore, the use of a critical load that 

would be associated with a highly 

sensitive water body in a naturally 

acidic system would impose a high 

degree of relative protection in terms of 

allowable ambient air concentrations of 

oxides of nitrogen and sulfur and 


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related deposition, while potentially 

affording little or no public welfare 

benefit from attempting to improve a 

naturally acidic system. 

Based on these considerations, the PA 

concludes it is appropriate to consider 

the use of a range of percentiles that 

extends lower than the range identified 

above for acid-sensitive ecoregions. 

Consideration of a lower percentile 

would avoid representing a relatively 

non-acid sensitive ecoregion by a 

critical load associated with relatively 

more acid-sensitive water bodies. In 

particular, the PA concludes it is 

appropriate to focus on the median or 

50th percentile of the distribution of 

critical loads so as to avoid overprotection in such ecoregions. 

Recognizing that relatively non-acid 

sensitive ecoregions generally are not 

sampled to the extent that acid-sensitive 

ecoregions are, it also is appropriate to 

consider using the median critical load 

of all relatively non-acid sensitive 

ecoregions for each such ecoregion. 

ii. Factor F2 

As discussed above, factor F2 is the 

amount of reduced nitrogen deposition 

within an ecoregion, including the 

deposition of both ammonia gas and 

ammonium ion, and is defined as: F2 = 

NHX/Qr. The PA calculated the 

representative runoff rate, Qr, using a 

similar approach as noted above for 

factor F1; i.e., the median value of the 

distribution of Q values that are 

available for water bodies within each 

ecoregion. In the PA, 2005 CMAQ 

model simulations over 12-km grids are 

used to calculate an average value of 

NHX for each ecoregion. The NHX term 

is based on annual average model 

outputs for each grid cell, which are 

spatially averaged across all the grid 

cells contained in each ecoregion to 

calculate a representative annual 

average value for each ecoregion. The 

PA concludes that this approach of 

using spatially averaged values is 

appropriate for modeling, largely due to 

the relatively rapid mixing of air masses 

that typically results in relatively 

homogeneous air quality patterns for 

regionally dispersed pollutants. In 

addition, there is greater confidence in 

using spatially averaged modeled 

atmospheric fields than in using 

modeled point-specific fields. 

This averaging approach is also used 

for the air concentration and deposition 

terms in factors F3 and F4, as discussed 

below. The PA notes that modeled NHX

deposition exhibits greater spatial 

variability than the other modeled terms 

in factors F3 and F4. Recognizing this 

greater variability, the PA concludes 

that it would be appropriate to consider 

alternative approaches to specifying the 

value of NHX. One such approach might 

involve the use of more localized and/ 

or contemporaneous modeling in areas 

where this term is likely to be 

particularly variable and important. 

iii. Factors F3 and F4 

As discussed above, factors F3 and F4 

are the ratios that relate ambient air 

concentrations of NOy and SOX to the 

associated deposition, and are defined 

as follow: F3 = TNOy/ Qr and F4 = TSOx/ 

Qr. TNOy is the transference ratio that 

converts ambient air concentrations of 

NOy to deposition of NOy and TSOx is 

the transference ratio that converts 

ambient air concentrations of SOX to 

deposition of SOX. The representative 

runoff rate, Qr, is calculated as for 

factors F1 and F2. The transference 

ratios are based on the 2005 CMAQ 

simulations, using average values for 

each ecoregion, as noted above for factor 

F2. More specifically, the transference 

ratios are calculated as the annual 

deposition of NOy or SOX spatially 

averaged across the ecoregion and 

divided by the annual ambient air 

concentration of NOy or SOX, 

respectively, spatially averaged across 

the ecoregion. 

c. Factors in Data-Limited Ecoregions 

As discussed above in section 

III.B.5.a, in the PA the initial 

delineation of acid-sensitive and 

relatively non-acid sensitive ecoregions 

was based on available ANC and 

alkalinity data. Areas not meeting the 

ANC criteria described above are 

categorized as relatively non-acid 

sensitive. The development of a 

reasonable distribution of critical loads 

for water bodies within an ecoregion for 

the purpose of identifying the 

representative critical load requires 

additional data, including more specific 

water quality data for major cations and 

anions. This means that the water 

bodies that can be used to develop a 

distribution of critical loads is generally 

a subset of those water bodies for which 

ANC data are available Consequently, 

there are certain ecoregions with sparse 

data that are not suitable for developing 

a distribution of critical loads. 

As noted above, the PA judges that it 

is not appropriate to develop such 

distributions based on data from less 

than ten water bodies within an 

ecoregion. Twelve such ecoregions, 

which included only relatively non-acid 

sensitive ecoregions, were characterized 

as being data-limited. For these 

ecoregions, the PA considered 

alternative approaches to specifying 

values for the terms CLr and Qr for the 

purpose of determining values for each 

of the factors in the AAI equation. For 

these data-limited ecoregions, the PA 

judges that it is appropriate to use the 

median values of CLr and Qr from the 

distributions of these terms for all other 

relatively non-acid sensitive ecoregions, 

rather than attempting to use severely 

limited data to develop a value for these 

terms based solely on data from such an 

ecoregion. Further, consideration could 

be given to using a single national 

default value for all relatively non-acid 

sensitive ecoregions. The PA notes that 

this data limitation is not a concern in 

specifying values for the other terms in 

the AAI equation for such ecoregions, 

since those terms are based on data from 

the 2005 CMAQ model simulation, 

which covers all ecoregions across the 

contiguous U.S. 

d. Application to Hawaii, Alaska, and 

the U.S. Territories 

The above methods for specifying 

ecoregion-specific values for the factors 

in the AAI equation apply to those 

ecoregions within the contiguous U.S. 

For areas outside the continental U.S., 

including Hawaii, Alaska, and the U.S. 

Territories, there is currently a lack of 

available data to characterize the 

sensitivity of such areas, as well as a 

lack of water body-specific data and 

CMAQ-type modeling to specify values 

for the F1 through F4 factors. Thus, the 

PA has considered possible alternative 

approaches to specifying values for 

factors F1 through F4 in the AAI 

equation for these areas. 

One such approach could be to 

specify area-specific values for the 

factors based on values derived for 

ecoregions with similar acid 

sensitivities, to the extent that relevant 

information can be obtained to 

determine such similarities. Such an 

approach would involve conducting an 

analysis to characterize similarities in 

relevant ecological attributes between 

ecoregions in the contiguous U.S. and 

these areas outside the contiguous U.S. 

so as to determine the appropriateness 

of utilizing ecoregion-specific values for 

the CLr and Qr terms from one or more 

ecoregions within the contiguous U.S. 

This approach would also involve 

conducting additional air quality 

modeling for the areas that are outside 

the geographical scope of the currently 

available CMAQ model simulations, so 

as to develop the other information 

necessary to specify values for factors 

F2 through F4 for these areas. 

A second approach could rely on 

future data collection efforts to establish 

relevant ecological data within these 

areas that, together with additional air 

quality modeling, could be used to 

specify area-specific values for factors 


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F1 through F4. Until such time as 

relevant data become available, these 

areas could be treated the same as datalimited ecoregions in the contiguous 

U.S. that are relatively non-acid 

sensitive. 

The PA concludes that either 

approach would introduce substantial 

uncertainties that arise from attempting 

to extrapolate values based on similarity 

assumptions or arbitrarily assigning 

values for factors in the AAI equation 

that would be applicable to these areas 

outside the contiguous U.S. In light of 

such uncertainties, the PA concludes 

that it would also be appropriate to 

consider relying on the existing NO2

and SO2 secondary standards in these 

areas for protection of any potential 

direct or deposition-related ecological 

effects that may be associated with the 

presence of oxides of nitrogen and 

sulfur in the ambient air. The PA 

concludes that relying on existing 

secondary standards in these areas is 

preferable to using a highly uncertain 

approach to allow for the application of 

a new standard based on the AAI in the 

absence of relevant area-specific data. 

6. Summary of the AAI Form 

With regard to the form of a multipollutant air quality standard to address 

deposition-related aquatic acidification 

effects, the PA concludes that 

consideration should be given to an 

ecologically relevant form that 

characterizes the relationships between 

the ambient air indicators for oxides of 

nitrogen and sulfur, the related 

deposition of nitrogen and sulfur, and 

the associated aquatic acidification 

effects in terms of a relevant ecological 

indicator. Based on the available 

information and assessments, 

consideration should be given to using 

ANC as the most appropriate ecological 

indicator for this purpose, in that it 

provides the most stable metric that is 

highly associated with the water quality 

properties that are directly responsible 

for the principal adverse effects 

associated with aquatic acidification: 

Fish mortality and reduced aquatic 

species diversity. 

The PA developed such a form, using 

a simple equation to calculate an AAI 

value in terms of the ambient air 

indicators of oxides and nitrogen and 

sulfur and the relevant ecological and 

atmospheric factors that modify the 

relationships between the ambient air 

indicators and ANC. Recognizing the 

spatial variability of such factors across 

the U.S., the PA concludes it is 

appropriate to divide the country into 

ecologically relevant regions, 

characterized as acid-sensitive or 

relatively non-acid-sensitive, and 

specify the value of each of the factors 

in the AAI equation for each such 

region. Omernik ecoregions, level III, are 

identified as the appropriate set of 

regions over which to define the AAI. 

There are 84 such ecoregions that cover 

the continental U.S. This set of 

ecoregions is based on grouping a 

variety of vegetation, geological, and 

hydrological attributes that are directly 

relevant to aquatic acidification 

assessments and that allow for a 

practical application of an aquatic 

acidification standard on a national 

scale. 

The PA defines AAI by the following 

equation: AAI = F1 ¥ F2 ¥ F3[NOy] ¥ 

F4[SOX]. Factors F1 through F4 would 

be defined for each ecoregion by 

specifying ecoregion-specific values for 

each factor based on monitored or 

modeled data that are representative of 

each ecoregion. The F1 factor is also 

defined by a target ANC value. More 

specifically: 

(1) F1 reflects a relative measure of an 

ecosystem’s ability to neutralize 

acidifying deposition. The value of F1 

for each ecoregion would be based on a 

representative critical load for the 

ecoregion associated with a single 

national target ANC level, as well as on 

a representative runoff rate. The 

representative runoff rate, which is also 

used in specifying values for the other 

factors, would be the median value of 

the distributions of runoff rates within 

the ecoregion. The representative 

critical load would be derived from a 

distribution of critical loads calculated 

for each water body in the ecoregion for 

which sufficient water quality and 

hydrology data are available. The 

representative critical load would be 

defined by selecting a specific 

percentile of the distribution. 

In identifying a range of percentiles 

that are appropriate to consider for this 

purpose, regions categorized as acid 

sensitive were considered separately 

from regions categorized as relatively 

non-acid sensitive. For acid sensitive 

regions, the PA concludes that 

consideration should be given to 

selecting a percentile from within the 

range of the 70th to the 90th percentile. 

The lower end of this range was selected 

to be appreciably above the median 

value so as to ensure that the critical 

load would be representative of the 

population of relatively more acid 

sensitive water bodies within the region, 

while the upper end was selected to 

avoid the use of a critical load from the 

extreme tail of the distribution which is 

subject to a high degree of variability 

and potential outliers. For relatively 

non-acid sensitive regions, the PA 

concludes that consideration should be 

given to selecting the 50th percentile to 

best represent the distribution of water 

bodies within such a region, or 

alternatively to using the median critical 

load of all relatively non-acid sensitive 

areas, recognizing that such areas are far 

less frequently evaluated than acid 

sensitive areas. Using either of these 

approaches would avoid characterizing 

a generally non-acid-sensitive region 

with a critical load that is representative 

of relatively acid sensitive water bodies 

that may exist within a generally nonacid sensitive region. 

(2) F2 reflects the deposition of 

reduced nitrogen. Consideration should 

be given to specifying the value of F2 for 

each region based on the averaged 

modeled value across the region, using 

national CMAQ modeling that has been 

conducted by EPA. Consideration could 

also be given to alternative approaches 

to specifying this value, such as the use 

of more localized and/or 

contemporaneous modeling in areas 

where this term is likely to be 

particularly variable and important. 

(3) F3 and F4 reflect transference 

ratios that convert ambient air 

concentrations of NOy and SOX, 

respectively, into related deposition of 

nitrogen and sulfur. Consideration 

should be given to specifying the values 

for F3 and F4 for each region based on 

CMAQ modeling results averaged across 

the region. We conclude that specifying 

the values or the transference ratios 

based on CMAQ modeling results alone 

is preferred to an alternative approach 

that combines CMAQ model estimates 

with observational data. 

(4) The terms [NOy] and [SOX] reflect 

ambient air concentrations measured at 

monitoring sites within each region. 

Using the equation, a value of AAI 

can be calculated for any measured 

values of ambient NOy and SOX. For 

such a NAAQS, the Administrator 

would set a single, national value for the 

level of the AAI used to determine 

achievement of the NAAQS, as 

discussed below in section III.D. The 

ecoregion-specific values for factors F1 

through F4 would be specified by EPA 

based on the most recent data and 

CMAQ model simulations, and codified 

as part of such a standard. These factors 

would be reviewed and updated as 

appropriate in the context of each 

periodic review of the NAAQS. 

The PA developed specific F factors 

for each ecoregion based on the 

approach discussed above, using 

alternative percentiles and alternative 

national target ANC levels. The results 

of this analysis for ecoregions 

characterized as acid sensitive are 

presented in Table 7–1a–d in the PA. 


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13Unlike 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. 

C. Averaging Time 

As discussed in section 7.3 of the PA, 

aquatic acidification can occur over 

both long- and short-term timescales. 

Long-term cumulative deposition of 

nitrogen and sulfur is reflected in the 

chronic acid-base balance of surface 

waters as indicated by measured annual 

ANC levels. Similarly, the use of steady 

state critical load modeling, which 

generates critical loads in terms of 

annual cumulative deposition of 

nitrogen and sulfur, means that the 

focus of ecological effects studies based 

on critical loads is on the long-term 

equilibrium status of water quality in 

aquatic ecosystems. Much of the 

evidence of adverse ecological effects 

associated with aquatic acidification, as 

discussed above in section II.A, is 

associated with chronically low ANC 

levels. Protection against a chronic ANC 

level that is too low is provided by 

reducing overall annual average 

deposition levels for nitrogen and 

sulfur. 

Reflecting this focus on long-term 

acidifying deposition, the PA developed 

the AAI that links ambient air indicators 

to deposition-related ecological effects, 

in terms of several factors, F1 through 

F4. As discussed above, these factors are 

all calculated as annual average values, 

whether based on water quality and 

hydrology data or on CMAQ model 

simulations. In the context of a standard 

defined in terms of the AAI, the PA 

concludes that it is appropriate to 

consider the same annual averaging 

time for the ambient air indicators as is 

used for the factors in the AAI equation. 

We also recognize that short-term (i.e., 

hours or days) episodic changes in 

water chemistry, often due to changes in 

the hydrologic flow paths, can have 

important biological effects in aquatic 

ecosystems. Such short-term changes in 

water chemistry are termed ‘‘episodic 

acidification.’’ Some streams may have 

chronic or base flow chemistry that is 

generally healthy for aquatic biota, but 

may be subject to occasional acidic 

episodes with potentially lethal 

consequences. Thus, short-term 

episodic ecological effects can occur 

even in the absence of long-term chronic 

acidification effects. 

Episodic declines in pH and ANC are 

nearly ubiquitous in drainage waters 

throughout the eastern U.S. Episodic 

acidification can result from several 

mechanisms related to changes in 

hydrologic flow paths. For example, 

snow can store nitrogen deposited 

throughout the winter and snowmelt 

can then release this stored nitrogen, 

together with nitrogen derived from 

nitrification in the soil itself, in a pulse 

that leads to episodic acidification in 

the absence of increased deposition 

during the actual episodic acidification 

event. The PA notes that inputs of 

nitrogen and sulfur from snowpack and 

atmospheric deposition largely cycle 

through soil. As a result, short-term 

direct deposition inputs are not 

necessarily important in episodic 

acidification. Thus, as noted in chapter 

3 of the ISA, protection against episodic 

acidity events can be achieved by 

establishing a higher chronic ANC level. 

Taken together, the above 

considerations support the conclusion 

that it is appropriate to consider the use 

of a long-term average for the ambient 

air indicators NOy and SOX for an 

aquatic acidification standard defined in 

terms of the AAI. The use of an annual 

averaging time for NOy and SOX

concentrations would be appropriate to 

provide protection against low chronic 

ANC levels, which in turn would 

protect against both long-term 

acidification and acute acidic episodes. 

The PA has also considered 

interannual variability in both ambient 

air quality and in precipitation, which 

is directly related to the deposition of 

oxides of nitrogen and sulfur from the 

ambient air. While ambient air 

concentrations show year-to-year 

variability, often the year-to-year 

variability in precipitation is 

considerably greater, given the highly 

stochastic nature of precipitation. The 

use of multiple years over which annual 

averages are determined would dampen 

the effects of interannual variability in 

both air quality and precipitation. For 

the ambient air indicators, the use of 

multiple-year averages would also add 

stability to calculations used to judge 

whether an area meets a standard 

defined in terms of the AAI. 

Consequently, the PA concludes that an 

annual averaging time based on the 

average of each year over a consecutive 

3- to 5-year period is appropriate to 

consider for the ambient air indicators 

NOy and SOX. In reaching this 

conclusion, the PA notes that in its 

comments on the second draft PA, 

CASAC agreed that a 3- to 5-year 

averaging time was appropriate to 

consider (Russell and Samet, 2010b). 

D. Level 

As discussed above, the PA concludes 

that ANC is the ecological indicator best 

suited to reflect the sensitivity of 

aquatic ecosystems to acidifying 

deposition from oxides of nitrogen and 

sulfur in the ambient air. The ANC is an 

indicator of the aquatic acidification 

expected to occur given the natural 

buffering capacity of an ecosystem and 

the loadings of nitrogen and sulfur 

resulting from atmospheric deposition. 

Thus, the PA developed a new standard 

for aquatic acidification that is based on 

the use of chronic ANC as the ecological 

indicator as a component in the AAI. 

The level of the standard would be 

defined in terms of a single, national 

value of the AAI. The standard would 

be met at a monitoring site when the 

multi-year average of the calculated 

annual values of the AAI was equal to 

or above the specified level of the 

standard.13 The annual values of the 

AAI would be calculated based on the 

AAI equation using the assigned 

ecoregion-specific values for factors F1 

through F4 and monitored annual 

average NOy and SOX concentrations. 

Since the AAI equation is based on 

chronic ANC as the ecological indicator, 

the level chosen for the standard would 

reflect a target chronic ANC value. As 

noted above, the assigned F factors for 

each ecoregion would be determined by 

EPA in the rulemaking to set the 

NAAQS, based on water quality and 

hydrology data, CMAQ modeling, the 

selected percentile that is used to 

identify a representative critical load 

within the ecoregion, and the selected 

level of the standard. The combination 

of the form of the standard, discussed 

above in section III.B, defined by the 

AAI equation and the assigned values of 

the F factors in the equation, other 

elements of the standard including the 

ambient air indicators (section III.A) and 

their averaging time (section III.C), and 

the level of the standard determines the 

allowable levels of NOy and SOX in the 

ambient air within each ecoregion. All 

of the elements of the standard together 

determine the degree of protection from 

adverse aquatic acidification effects 

associated with oxides of nitrogen and 

sulfur in the ambient air. The level of 

the standard plays a central role in 

determining the degree of protection 

provided and is discussed below. 

The PA focuses primarily on 

information that relates degrees of 

biological impairment associated with 

adverse ecological effects to aquatic 

ecosystems to alternative levels of ANC 

in reaching conclusions regarding the 

range of target ANC levels that is 

appropriate to consider for the level of 

the standard. The PA develops the 

rationale for identifying a range of target 

ANC levels that is appropriate to 

consider by addressing questions related 

to the following areas: (1) Associations 

between ANC and pH levels to provide 

an initial bounding for the range of ANC 


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values to be considered; (2) evidence 

that allows for the delineation of 

specific ANC ranges associated with 

varying degrees of severity of biological 

impairment ecological effects; (3) the 

role of ANC in affording protection 

against episodic acidity; (4) implications 

of the time lag response of ANC to 

changes in deposition; (5) past and 

current examples of target ANC values 

applied in environmental management 

practices; and (6) data linking public 

welfare benefits and ANC. 

1. Association Between pH Levels and 

Target ANC Levels 

As discussed above in section II.A 

and more fully in chapter 3 of the PA, 

specific levels of ANC are associated 

with differing levels of risk of biological 

impairment in aquatic ecosystems, with 

higher levels of ANC resulting in lower 

risk of ecosystem impacts, and lower 

ANC levels resulting in risk of both 

higher intensity of impacts and a 

broader set of impacts. While ANC is 

not the causal agent determining 

biological effects in aquatic ecosystem, 

it is a useful metric for determining the 

level at which a water body is protected 

against risks of acidification. There is a 

direct correlation between ANC and pH 

levels which, along with dissolved 

aluminum, are more closely linked to 

the biological causes of ecosystem 

response to acidification. 

Because there is a direct correlation 

between ANC and pH levels, the 

selection of target ANC levels is 

informed in part through information on 

effects of pH as well as direct studies of 

effects related to ANC. Levels of pH are 

closely associated with ANC in the pH 

range of approximately 4.5 to 7. Within 

this range, higher ANC levels are 

associated with higher pH levels. At a 

pH level of approximately 4.5, further 

reductions in ANC generally do not 

correlate with pH, as pH levels remain 

at approximately 4.5 while ANC values 

fall substantially. Similarly, at a pH 

value of approximately 7, ANC values 

can continue to increase with no 

corresponding increase in pH. As pH is 

the primary causal indicator of effects 

related to aquatic acidification, this 

suggests that ANC values below 

approximately ¥50 μeq/L (the apparent 

point in the relationship between pH 

and ANC where pH reaches a minimum) 

are not likely to result in further 

damage. In addition, ANC values 

around and above approximately 100 

μeq/L (the apparent region in the 

relationship where pH reaches a 

maximum) are not likely to confer 

additional protection. As a result, the 

initial focus in the PA was on target 

ANC values in the range of ¥50 to 100 

μeq/L. 

2. ANC Levels Related to Effects on 

Aquatic Ecosystems 

As discussed above in section II.A, 

the number of fish species present in a 

water body has been shown to be 

positively correlated with the ANC level 

in the water, with higher values 

supporting a greater richness and 

diversity of fish species. The diversity 

and distribution of phyto-zooplankton 

communities also are positively 

correlated with ANC. 

A summary of effects related to ANC 

ranges is shown above in Table II–1. 

Within the ANC range from 

approximately ¥50 to 100 μeq/L, linear 

and sigmoidal relationships are 

observed between ANC and ecosystem 

effects. On average, fish species richness 

is lower by one fish species for every 21 

μeq/L decrease in ANC in Shenandoah 

National Park streams (ISA, section 

3.2.3.4). As shown in Table II–1, ANC 

levels have been grouped into five 

categories related to expected ecological 

effects, including categories of acute 

concern (<0 μeq/L), severe concern (0– 

20 μeq/L), elevated concern (20–50 μeq/ 

L), moderate concern (50–100 μeq/L), 

and low concern (>100 μeq/L). This 

categorization is supported by a large 

body of research completed throughout 

the eastern U.S. (Sullivan et al., 2006). 

Water bodies with ANC values less 

than or equal to 0 μeq/L at based flow 

are chronically acidic. Such ANC levels 

can lead to complete loss of species and 

major changes in the ability of water 

bodies to support diverse biota, 

especially in water bodies that are 

highly sensitive to episodic 

acidification. Based on the above 

considerations, the PA has focused on 

target ANC levels no lower than 0 μeq/ 

L. 

As discussed in the PA, biota 

generally are not harmed when ANC 

values are >100 μeq/L, due to the low 

probability that pH levels will be below 

7. In the Adirondacks, the number of 

fish species also peaks at ANC values 

>100 μeq/L. This suggests that at ANC 

levels greater than 100 μeq/L, little risk 

from acidification exists in many 

aquatic ecosystems. At ANC levels 

below 100 μeq/L, overall health of 

aquatic communities can be maintained, 

although fish fitness and community 

diversity begin to decline. At ANC 

levels ranging from 100 down to 50 μeq/ 

L, there is increasing likelihood that the 

fitness of sensitive species (e.g., brook 

trout, zooplankton) will begin to 

decline. When ANC concentrations are 

below 50 μeq/L, the probability of 

acidification increases substantially, 

and negative effects on aquatic biota are 

observed, including large reductions in 

diversity of fish species and changes in 

the health of fish populations, affecting 

reproductive ability and fitness, 

especially in water bodies that are 

affected by episodic acidification. While 

there is evidence that ANC levels above 

50 can confer additional protection from 

adverse ecological effects associated 

with aquatic acidification in some 

sensitive ecosystems, the expectation 

that such incremental protection from 

adverse effects will continue up to an 

ANC level of 100 is substantially 

reduced. The PA concludes that the 

above considerations support a focus on 

target ANC levels up to a level greater 

than 50 μeq/L but below 100 μeq/L, 

such as up to a level of 75 μeq/L. 

In considering the available scientific 

evidence, as summarized here and 

discussed in more detail in the ISA and 

REA, in its review of the second draft 

PA, CASAC expressed the following 

views about the range of biological 

responses that corresponds to this range 

of ANC levels (i.e., 0–100 μeq/L): 

There will likely be biological effects of 

acidification at higher ANC values within 

this range, and there are relatively insensitive 

organisms that are not impacted at ANC 

values at the low end of this range. Adverse 

effects of acidification on aquatic biota are 

fairly certain at the low end of this range of 

ANC and incremental benefits of shifting 

waters to higher ANC become more uncertain 

at higher ANC levels. There is substantial 

confidence that there are adverse effects at 

ANC levels below 20 μeq/L, and reasonable 

confidence that there are adverse effects 

below 50 μeq/L. Levels of 50 μeq/L and 

higher would provide additional protection, 

but the Panel has less confidence in the 

significance of the incremental benefits as the 

level increases above 50 μeq/L. (Russell and 

Samet, 2010b) 

The PA concludes that the above 

considerations, including the views of 

CASAC, provide support for focusing on 

target ANC levels in the range of 20 to 

75 μeq/L. 

3. Consideration of Episodic Acidity 

As discussed in the PA, across the 

broad range of ANC values from 0 to 100 

μeq/L, ANC affords protection against 

the likelihood of decreased pH (and 

associated increases in Al) during long 

or short periods. In general, the higher 

the ANC within this range, the lower the 

probability of reaching low pH levels 

where direct effects such as increased 

fish mortality occur, as shown in Table 

3–1 of the PA. Accordingly, greater 

protection would be achieved by target 

chronic ANC values set high enough to 

avoid pH depression to levels associated 

with elevated risk. 


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The specific relationship between 

ANC and the probability of reaching pH 

levels of elevated risk varies by water 

body and fish species. The ANC levels 

below 20 μeq/L are generally associated 

with high probability of low pH, leading 

to death or loss of fitness of biota that 

are sensitive to acidification (US EPA, 

2008, section 5.2.2.1; US EPA, 2009, 

section 5.2.1.2). At these levels, during 

episodes of high acidifying deposition, 

brook trout populations may experience 

lethal effects. In addition, the diversity 

and distribution of zooplankton 

communities decline sharply at ANC 

levels below 20 μeq/L. Overall, there is 

little uncertainty that significant effects 

on aquatic biota are occurring at ANC 

levels below 20 μeq/L. 

It is clear that at ANC levels 

approaching 0 μeq/L (Table II–1), there 

is significant impairment of sensitive 

aquatic ecosystems with almost 

complete loss of fish species. Avoiding 

ANC levels approaching 0 μeq/L is 

particularly relevant to episodic spikes 

in acidity that occur during periods of 

rapid snow melt and during and after 

major precipitation events. Since the 

ANC range considered in the PA reflects 

average, long-term base flow values, it is 

appropriate to consider protecting 

against episodic drops in ANC values to 

a level as low as 0 μeq/L. Staddard et 

al. (2003) noted on average a 30 μeq/L 

depression of ANC between spring and 

summer time values, indicating the 

need to maintain higher base flow ANC 

levels to protect against ANC levels 

below 0 μeq/L. The above 

considerations do not provide support 

for a target chronic ANC level as low as 

0 μeq/L for a standard that would 

protect against significant harm to 

aquatic ecosystems, including harm 

from episodic acidification. The PA 

concludes that these considerations also 

support a lower end of the range for 

consideration no lower than 20 μeq/L. 

The CASAC agreed with this 

conclusion in its comments on the 

second draft PA (Russell and Samet, 

2010b). The CASAC noted that ‘‘there 

are clear and marked biological effects 

at ANC values near 0 μeq/L, so this is 

probably not an appropriate target 

value’’ for the AAI. With regard to the 

likelihood of impairment of aquatic 

ecosystems due to episodic 

acidification, in terms of specific target 

levels for chronic ANC, CASAC 

expressed the following view: 

Based on surface waters studied in the 

Northeast, decreases in ANC associated with 

snowmelt [are] approximately 50 μeq/L. 

Thus, based on these studies, a long term 

ANC target level of 75 μeq/L would generally 

guard against effects from episodic 

acidification down to a level of about 25 μeq/ 

L. (Russell and Samet, 2010b) 

4. Consideration of Ecosystem Response 

Time 

The PA notes that when considering 

a standard level to protect against 

aquatic acidification, it is appropriate to 

take into account both the time period 

to recovery as well as the potential for 

recovery in acid-sensitive ecoregions. 

Ecosystems become adversely impacted 

by acidifying deposition over long 

periods of time and have variable time 

frames and abilities to recover from 

such perturbations. Modeling presented 

in the REA (U.S. EPA, 2009, section 

4.2.4) shows the estimated ANC values 

for Adirondack lakes and Shenandoah 

streams under pre-acidification 

conditions and indicates that for a small 

percentage of lakes and streams, natural 

ANC levels would have been below 50 

μeq/L. Therefore, for these water bodies, 

reductions in acidifying deposition are 

not likely to achieve an ANC of 50 μeq/ 

L or greater. Conversely, for some lakes 

and streams the level of perturbation 

from long periods of acidifying 

deposition has resulted in very low 

ANC values compared to estimated 

natural conditions. For such water 

bodies, the time to recovery would be 

largely dependent on future inputs of 

acidifying deposition. 

Setting a standard level in terms of a 

target chronic ANC level is based on the 

long-term response of aquatic 

ecosystems. The time required for a 

water body to achieve the target ANC 

level—given a decrease in ambient air 

concentrations of NOy and SOx and 

related acidifying deposition such that 

the critical load for a target ANC is not 

exceeded—is often decades if not 

centuries. In recognition of the potential 

public welfare benefits of achieving the 

target ANC in a shorter time frame, the 

concept of target loads had been 

developed. Target loads represent the 

depositional loading that is expected to 

achieve a particular level of the 

ecological indicator by a given time. For 

example, to achieve an ANC level of 20 

μeq/L by 2030, it might be necessary to 

specify a higher target ANC level of, for 

example, 50 μeq/L, such that the 

depositional loading would be reduced 

more quickly than would occur if the 

depositional loading was based on 

achieving a target ANC level of 20 μeq/ 

L as a long-term equilibrium level. In 

this example, the target ANC of 50 μeq/ 

L would ultimately be realized many 

years later. 

The above considerations have 

implications for selecting an appropriate 

standard level, in that the standard level 

affects not only the ultimate degree of 

protection that would be afforded by the 

standard, but also the time frame in 

which such protection would be 

realized. However, the PA recognizes 

that there is a great deal of heterogeneity 

in response times among water bodies 

and that there is only very limited 

information from dynamic modeling 

that would help to quantify recovery 

time frames in areas across the country. 

As a consequence, quantification of a 

general relationship between critical 

loads associated with a specific longterm target ANC level and target loads 

associated with achieving the target 

ANC level within a specific time frame 

is not currently possible. Thus, while 

the time frame for recovery is an 

important consideration in selecting an 

appropriate range of levels to consider, 

the PA concludes that it can only be 

considered in a qualitative sense at this 

time. 

5. Prior Examples of Target ANC Levels 

A number of regional organizations, 

states, and international organizations 

have developed critical load frameworks 

to protect against acidification of 

sensitive aquatic ecosystems. In 

considering the appropriate range of 

target ANC levels for consideration in 

this review, it is informative to evaluate 

the target ANC levels selected by these 

different organizations, as well as the 

rationale provided in support of the 

selected levels. Chapter 4 of the PA 

provides a detailed discussion of how 

critical loads have been developed and 

used in other contexts. Specific target 

values and their rationales are 

summarized below. 

The UNECE has developed critical 

loads in support of international 

emissions reduction agreements. As 

noted in chapter 4 of the PA, critical 

loads were established to protect 95 

percent of surface waters in Europe from 

an ANC less than 20 μeq/L based on 

protection of brown trout. Individual 

countries have set alternative ANC 

targets; for example, Norway targets an 

ANC of 30 μeq/L based on protection of 

Atlantic salmon. Several states have 

established target ANC or pH values 

related to protection of lakes and 

streams from acidification. While 

recognizing that some lakes in the 

Adirondacks will have a naturally low 

pH, the state of New York has 

established a target pH value of 6.5 for 

lakes that are not naturally below 6.5. 

As noted above, this level is associated 

with an ANC value that is likely to be 

between 20 and 50 μeq/L or possibly 

higher. New Hampshire and Vermont 

have set ANC targets of 60 μeq/L and 50 

μeq/L, respectively. Tennessee has 

established site-specific target ANC 


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values based on assessments of natural 

acidity, with a default value of 50 μeq/ 

L when specific data are not available. 

Taken together, these policy 

responses to concerns about ecological 

effects associated with aquatic 

acidification indicate that target ANC 

values between 20 and 60 μeq/L have 

been selected by states and other 

nations to provide protection of lakes 

and streams in some of the more 

sensitive aquatic ecosystems. 

6. Consideration of Public Welfare 

Benefits 

The point at which effects on public 

welfare become adverse is not defined 

in the CAA. Characterizing a known or 

anticipated adverse effect to public 

welfare is an important component of 

developing any secondary NAAQS. 

According to the CAA, welfare effects 

include: 

* * * effects on soils, water, crops, 

vegetation, manmade materials, animals, 

wildlife, weather, visibility, and climate, 

damage to and deterioration of property, and 

hazards to transportation, as well as effect on 

economic values and on personal comfort 

and well-being, whether caused by 

transformation, conversion, or combination 

with other air pollutants. (CAA, section 

302(h)). 

Consideration of adversity to public 

welfare in the context of the secondary 

NAAQS for oxides of nitrogen and 

sulfur can be informed by information 

about losses in ecosystem services 

associated with acidifying deposition 

and the potential economic value of 

those losses, as summarized above in 

section II.C and discussed more fully in 

chapter 4 of the PA. 

Ecosystem service losses at alternative 

ANC levels are difficult to enumerate. 

However, in general there are categories 

of ecosystem services, discussed in 

chapter 4 of the PA, that are related to 

the specific ecosystem damages 

expected to occur at alternative ANC 

levels. Losses in fish populations due to 

very low ANC (below 20 μeq/L) are 

likely associated with significant losses 

in value for recreational and subsistence 

fishers. Many acid sensitive lakes are 

located in areas with high levels of 

recreational fishing activity. For 

example, in the northeastern U.S., 

where nearly 8 percent of lakes are 

considered acidic, more than 9 percent 

of adults participate in freshwater 

fishing, with an estimated value of 

approximately $5 billion in 2006. This 

suggests that improvements in lake fish 

populations may be associated with 

significant recreational fishing value. 

As discussed in the PA, inland 

surface waters also provide cultural 

services such as aesthetic and existence 

value and educational services. To the 

extent that piscivorous birds and other 

wildlife are harmed by the absence of 

fish in these waters, hunting and 

birdwatching activities are likely to be 

adversely affected. A case study of the 

value to New York residents of 

improving the health of lakes in the 

Adirondacks found significant 

willingness to pay for those 

improvements. When scaled to evaluate 

the improvement in lake health from 

achieving ANC values of either 20 or 50 

μeq/L, the study implies benefits to the 

New York population roughly on the 

order of $300–900 million per year (in 

constant 2007$). The survey 

administered in this study recognized 

that participants were thinking about 

the full range of services provided by 

the lakes in question—not just the 

recreational fishing services. Therefore 

the estimates of willingness to pay 

include resident’s benefits for potential 

hunting and birdwatching activities and 

other ancillary services. These results 

are just for New York populations. The 

PA concludes that if similar benefits 

exist for improvements in other acid 

sensitive lakes, the economic value to 

U.S. populations could be very 

substantial, suggesting that, at least by 

one measure of impact on public 

welfare, impacts associated with ANC 

less than 50 μeq/L may be adverse to 

public welfare. 

7. Summary of Alternative Levels 

Based on all the above considerations, 

the PA concludes that consideration 

should be given to a range of standard 

levels from 20 to 75 μeq/L. The available 

evidence indicates that target ANC 

levels below 20 μeq/L would be 

inadequate to protect against substantial 

ecological effects and potential 

catastrophic loss of ecosystem function 

in some sensitive aquatic ecosystems. 

While ecological effects occur at ANC 

levels below 50 μeq/L in some sensitive 

ecosystems, the degree and nature of 

those effects are less significant than at 

levels below 20 μeq/L. Levels at and 

above 50 μeq/L would be expected to 

provide additional protection, although 

uncertainties regarding the potential for 

additional protection from adverse 

ecological effects are much larger for 

target ANC levels above about 75 μeq/ 

L, as effects are generally appreciably 

less sensitive to changes in ANC at such 

higher levels. 

In reaching this conclusion in the PA, 

consideration was given to the extent to 

which a target ANC level within this 

range would protect against episodic as 

well as long-term ecological effects. 

Levels in the mid- to upper-part of this 

range would be expected to provide 

greater protection against short-term, 

episodic peaks in aquatic acidification, 

while lower levels within this range 

would give more weight to protection 

from long-term rather than episodic 

acidification. Similarly, levels in the 

mid- to upper-part of this range would 

be expected to result in shorter time 

periods for recovery given the lag in 

ecosystem response in some sensitive 

ecosystems relative to levels in the 

lower part of this range. The PA also 

notes that this range encompasses target 

ANC values that have been established 

by various States and regional and 

international organizations to protect 

against acidification of aquatic 

ecosystems. 

The PA recognizes that the level of 

the standard together with the other 

elements of the standard, including the 

ambient air indicators, averaging time, 

and form, determine the overall 

protectiveness of the standard. Thus, 

consideration of a standard level should 

reflect the strengths and limitations of 

the evidence and assessments as well as 

the inherent uncertainties in the 

development of each of the elements of 

the standard. The implications of 

considering alternative standards, 

defined in terms of alternative 

combinations of levels and percentile 

values that are a critical component of 

factor F1 in the form of the standard, are 

discussed below in section III.E. Key 

uncertainties in the various components 

of the standard are summarized and 

considered below in section III.F. 

E. Combined Alternative Levels and 

Forms 

To provide some perspective on the 

implications of various alternative 

multi-pollutant, AAI-based standards, 

the PA presented the number of acidsensitive ecoregions that would likely 

not meet various sets of alternative 

standards. The alternative standards 

considered were based on combinations 

of alternative target ANC levels, within 

the range of 20 to 75 μeq/L, and 

alternative forms, characterized by 

alternative representative percentiles 

within the range of the 70th to 90th 

percentile. These alternative standards 

are also defined in terms of the other 

elements of the standard: ambient air 

indicators NOy and SOx, discussed 

above in section III.A; other elements of 

the form of the standard, including 

ecoregion-specific values for factors F1 

through F4 in the AAI equation, 

discussed above in section III.B.5; and 

an annual averaging time for NOy and 

SOx, discussed above in section III.C. 

With regard to the averaging time, the 

assessment did not consider multi-year 

averaging of the calculated annual AAI 


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14Tables 7–1a–d and 7–2 in the PA present 

assessment results for 29 ecoregions that had been 

initially characterized as acid sensitive. 

Subsequently, based on a broader set of criteria 

used to characterize ecoregions as acid sensitive, as 

discussed above in section III.B.5.a, the set of 

ecoregions characterized as acid sensitive was 

narrowed to include 22 ecoregions. 

values due to data limitations, 

including, for example, the lack of 

CMAQ modeling for multiple 

consecutive years. In this assessment, 

we characterize an ecoregion as likely 

not meeting a given alternative standard 

if the calculated AAI value is less than 

the target ANC level of the standard, 

recognizing that higher AAI values are 

more protective than lower values. 

The results of this assessment are 

presented in Table 7–1a–d in the PA for 

a subset of ecoregions including those 

characterized as acid sensitive. 

Calculated annual AAI values at the 

ecoregion level are shown for each 

alternative standard considered. Based 

on these AAI values, Table 7–2 in the 

PA summarizes the number of acidsensitive ecoregions that would likely 

not meet each of the alternative 

standards considered.14 Calculated AAI 

values for all ecoregions categorized as 

relatively non-acid sensitive are shown 

in Table D–5 in Appendix D of the PA. 

In all cases, these relatively non-acid 

sensitive ecoregions were estimated to 

meet all of the alternative standards 

considered in this assessment. 

As described above, the AAI values 

presented in Table 7–1a–d of the PA are 

based in part on data from 2005 CMAQ 

model simulations, which was used to 

generate values for F2 through F4 in the 

AAI equation, as well as to estimate 

annual average ambient air 

concentrations of NOy and SOx that 

reflect recent air quality in the absence 

of currently available monitored 

concentrations in sensitive ecoregions 

across the country. Water quality and 

hydrology data from water bodies 

within each ecoregion were also used in 

calculating the AAI values. Such data 

were initially used to calculate critical 

loads for each water body with 

sufficient data within an ecoregion so as 

to identify the nth percentile critical 

load representative of the ecoregion 

used in calculating the F1 factor for the 

ecoregion. As expected, the number of 

ecoregions that likely would not meet 

alternative standards increases with 

increasing percentile values and target 

ANC levels (U.S. EPA, 2011, Table 7–2). 

Out of 22 acid-sensitive ecoregions, the 

number of ecoregions that would likely 

not meet the alternative standards 

ranges from 22 for the most protective 

alternative standard considered (75 μeq/ 

L, 90th percentile) to 4 for the least 

protective alternative standard (20 μeq/ 

L, 70th percentile). It is apparent that 

both the percentile and the level chosen 

have a strong influence, over the ranges 

considered, in determining the number 

of areas that would likely not meet this 

set of alternative standards. 

The PA observes that there is one 

grouping of these acid-sensitive 

ecoregions that would likely not meet 

almost all combinations of level and 

form under consideration (U.S. EPA, 

2011, Table 7–2 and Appendix D). This 

group is made up of southern 

Appalachian mountain areas, including 

North Central Appalachians, 5.3.3; 

Ridge and Valley, 8.4.1; Central 

Appalachians, 8.4.2; Blue Ridge, 8.4.4; 

and Southwestern Appalachians, 8.4.9. 

In addition, these ecoregions exhibit the 

highest amounts of exceedance relative 

to alternative standards. 

The Northern Appalachian and 

Atlantic Maritime Highlands (5.3.1), 

which includes the Adirondacks, and 

the Northern Lakes and Forests (5.2.1) of 

the upper midwest exhibit similar 

patterns with respect to in the role of 

level and percentile in identifying 

regions not likely to meet alternative 

standards, although there are 

considerably fewer cases compared to 

the regions in the Appalachians. 

In the mountainous west, the Sierra 

Nevada (6.2.12), Idaho Batholith (6.2.15) 

and the Cascades (6.2.7) ecoregions 

likely would not meet alternative 

standards in fewer cases relative to 

eastern regions, with the Sierra Nevada 

ecoregion exhibiting relatively greater 

sensitivity compared to all western 

regions. Only in the upper part of the 

ranges of level and percentile do regions 

in the northern and central Rockies 

likely not meet alternative standards. 

In considering these findings, the PA 

observes that the standard as defined by 

the AAI behaves in an intuitively logical 

manner. That is, an increase in 

ecoregions likely not to meet the 

standard is associated with higher 

alternative levels and percentiles, both 

of which contribute to a lower 

regionally representative critical load. 

Moreover, the areas of known adverse 

aquatic acidification effects are 

identified, mostly in high elevation 

regions or in the northern latitudes—the 

Adirondacks, Shenandoahs, northern 

midwest lakes and the mountainous 

west. These results reflect the first 

application of a nationwide model that 

integrates water quality and 

atmospheric processes at a national 

scale and provides findings that are 

consistent with our basic understanding 

of the extent of aquatic acidification 

across the U.S. What is particularly 

noteworthy is that this model is not 

initialized with a starting ANC based on 

water quality data, which likely would 

result in a reproduction of water quality 

observations. Rather, this standard 

reflects the potential of the changes in 

atmospheric concentrations of NOy and 

SOx to induce long-term sustained 

changes in surface water systems. The 

PA notes that the fact that the patterns 

of adversity based on applying this 

standard are commensurate with what is 

observed in surface water systems 

provides confidence in the basic 

underlying formulation of the standard. 

The PA notes that the Appalachian 

mountain regions merit further 

inspection as they stand out as areas 

with the largest relative exceedances 

from a national perspective. Water 

quality data from these regions as well 

as an emissions sensitivity CMAQ 

simulation were considered to better 

understand the simulated behavior of 

these regions. The maps and tables in 

appendix D of the PA include paired 

comparisons of the CMAQ 2005 and 

emissions sensitivity simulations. The 

emissions sensitivity simulation reflects 

domain-wide reductions in NOy and 

SOx emissions of 48 percent and 42 

percent, respectively, relative to 2005 

base year emissions. The PA assumes 

that this emissions sensitivity 

simulation is indicative of future 

conditions. 

The emissions sensitivity results 

project that many of the regions that 

likely would not meet the alternative 

standards based on recent air quality, 

especially at alternative levels of 20 and 

35 μeq/L, would likely meet such 

standards in the future year scenario for 

the Appalachian mountain regions. It is 

apparent that the AAI calculations are 

especially sensitive to changes in SOx

emissions as the Appalachian regions 

have the highest SOx concentrations and 

deposition rates (U.S. EPA, 2011,section 

2), and the AAI equation responds as 

expected to modeled reductions in SOx. 

The emissions sensitivity scenario is a 

prospective application of the standard, 

in the sense that rules derived from the 

air quality management process result in 

reductions of NOy and SOx emissions. 

Expected emission changes over the 

next two decades should be far greater 

than the 42 percent and 48 percent, 

respectively, SOx and NOy reductions 

used in this analysis, with a consequent 

further reduction in areas that would 

likely not meet alternative standards. 

The Appalachian mountain regions 

generally have low DOC levels, average 

runoff rates, moderately low base cation 

supply and highly elevated sulfate 

concentrations. Collectively, those 

attributes do not suggest naturally acidic 

conditions as the availability of 


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anthropogenic contributions of mineral 

acids is likely responsible for observed 

low ANC values in those regions. 

The PA notes the Sierra Nevada 

region as an interesting case study, as it 

has some of the lowest critical load 

values nationally (U.S. EPA, 2011, Table 

D–3). Water quality data indicate 

extremely low sulfate, as expected given 

the relatively low SO2 emissions in the 

western U.S. Extremely low base cation 

supply and low Neco, which mitigate 

the effect of nitrogen deposition, explain 

the low critical load values. Low Neco 

values appear to associate well with 

high elevation western U.S. regions, 

perhaps reflecting the more arid and 

reduced vegetation density relative to 

eastern U.S. regions. The proximity to 

high level nitrogen emissions combined 

with very low base cation supply 

explains the cases where the Sierra 

region likely does not meet alternative 

standards. Because Neco values are low 

in the Sierras, the system responds 

effectively to reductions of NOx

emissions, as illustrated in the maps 

and tables of Appendix D of the PA. 

Although Neco affords protection from 

the acidifying effects of nitrogen 

deposition, the availability of excessive 

nitrogen neutralization capacity also 

means that reductions in nitrogen are 

not as effective as reductions in SOx in 

reducing the calculated AAI. 

In reviewing these results, the PA 

observes that the analysis of the 

alternative combinations of level and 

form presented provide context for 

considering the impact of different 

standards. Since the AAI equation has 

been newly developed in the PA, these 

examples of estimated exceedances help 

to address the question of whether the 

AAI equation responds in a reasonable 

manner with regard to identifying areas 

of concern and to prospective changes 

in atmospheric conditions likely to 

result from future emissions reduction 

strategies. The PA concludes that the 

behavior of the AAI calculations is both 

reasonable and explainable, which the 

PA concludes serves to increase 

confidence in considering a standard 

defined in terms of the AAI. 

F. Characterization of Uncertainties 

This section summarizes discussions 

of the results of analyses and 

assessments, presented more fully in the 

PA (U.S. EPA, 2011, section 7.6 and 

Appendices F and G), intended to 

address the relative confidence 

associated with the linked atmosphericecological effects system described 

above. An overview of uncertainties is 

presented in the context of the major 

structural components underlying the 

standard, as well as with regard to areas 

of relatively high uncertainty. The 

section closes with a discussion of data 

gaps and uncertainties associated with 

the use of ecological and atmospheric 

modeling to specify the factors in the 

AAI equation, which can be used to 

guide future field programs and longerterm research efforts. 

1. Overview of Uncertainty 

As discussed in the PA (U.S. EPA, 

2011, Table 7–3), there is relatively low 

uncertainty with regard to the 

conceptual formulation of the overall 

structure of the AAI-based standard that 

incorporates the major associations 

linking biological effects to air 

concentrations. Based on the strength of 

the evidence that links species richness 

and mortality to water quality, the 

associations are strongly causal and 

without any obvious confounding 

influence. The strong association 

between the ecosystem indicator (ANC) 

and the causative water chemistry 

species (dissolved aluminum and 

hydrogen ion) reinforces the confidence 

in the linkage between deposition of 

nitrogen and sulfur and effects. This 

strong association between ANC and 

effects is supported by a sound 

mechanistic foundation between 

deposition and ANC. The same 

mechanistic strength holds true for the 

relationship between ambient air levels 

of nitrogen and sulfur and deposition, 

which completes the linkage from 

ambient air indicators through 

deposition to ecological effects. 

There are relatively higher 

uncertainties, however, in considering 

specific elements within the structure of 

an AAI-based standard, including the 

deposition of SOx, NOy, and NHx as well 

as the critical load-related component, 

each of which can vary within and 

across ecoregions. Overall system 

uncertainty relates not just to the 

uncertainty in each such element, but 

also to the combined uncertainties that 

result from linking these elements 

together within the AAI-based structure. 

Some of these elements—including, for 

example, dry deposition, pre-industrial 

base cation production, and reduced 

nitrogen deposition—are estimated with 

less confidence than other elements 

(U.S. EPA, 2011, Table 7.3). The 

uncertainties associated with all of these 

elements, and the combination of these 

elements through the AAI equation, are 

discussed below and in the following 

sections related to measured data gaps 

and modeled processes for both air 

quality and water quality. 

The lack of observed dry deposition 

data is constrained by resources and the 

lack of efficient measurement 

technologies. Progress in reducing 

uncertainties in dry deposition will 

depend on improved atmospheric 

concentration data and direct deposition 

flux measurements of the relevant suite 

of NOy and SOx species. 

Pre-industrial base cation 

productivity by definition is not 

observable. Contemporary observations 

and inter-model comparisons are useful 

tools that would help reduce the 

uncertainty in estimates of preindustrial 

base cation productivity used in the AAI 

equation. In characterizing 

contemporary base cation flux using 

basic water quality measurements (i.e., 

major anion and cation species as 

defined in equation 2.11 in the PA), it 

is reasonable to assume that a major 

component of contemporary base cation 

flux is associated with pre-industrial 

weathering rates. To the extent that 

multiple models converge on similar 

solutions, greater confidence in 

estimating pre-industrial base cation 

production would be achieved. 

Characterization of NHx deposition 

has been evolving over the last decade. 

The relatively high uncertainty in 

characterizing NHx deposition is due to 

both the lack of field measurements and 

the inherent complexity of 

characterizing NHx with respect to 

source emissions and dry deposition. 

Because ammonia emissions are 

generated through a combination of 

man-made and biological activities, and 

ammonia is semi-volatile, the ability to 

characterize spatial and temporal 

distributions of NHx concentrations and 

deposition patterns is challenging. 

While direct measurement of NHx

deposition is resource intensive because 

of the diffuse nature of sources (i.e., 

area-wide and non-point sources), there 

have been more frequent deposition flux 

studies, relative to other nitrogen 

species, that enable the estimation of 

both emissions and dry deposition. 

Also, while ammonia has a relatively 

high deposition velocity and 

traditionally was thought to deposit 

close to the emissions release areas, the 

semi-volatile nature of ammonia results 

in re-entrainment back into the lower 

boundary layer resulting in a more 

dispersed concentration pattern 

exhibiting transport type characteristics 

similar to longer lived atmospheric 

species. These inherent complexities in 

source characterization and ambient 

concentration patterns raise the 

uncertainty level of NHx in general. 

However, the PA notes that progress is 

being made in measuring ammonia with 

cost efficient samplers and anticipates 

the gradual evolution of a spatially 

robust ammonia sampling network that 

would help support analyses to reduce 

underlying uncertainties in NHx


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deposition. Also, from an aquatic 

acidification perspective, NHx is not as 

important a driver as NOy and SOx in 

the mountainous areas in the eastern 

U.S. However, the relative importance 

of NHx is likely to increase over time, 

in light of air quality rules in place 

designed to reduce emissions of NOy

and SOx. 

2. Uncertainties Associated With Data 

Gaps 

In summarizing uncertainties with 

respect to available measurement data 

and the use of ecological and 

atmospheric models, the PA indentified 

data gaps and model uncertainties in 

relative terms by comparing, for 

example, the relative richness of data 

between geographic areas or 

environmental media. With regard to 

relevant air quality measurements, the 

PA notes that such measurements are 

relatively sparse in the western U.S. 

While the spatial extent of CASTNET 

coverage has gradually incorporated 

western U.S. locations with support 

from the NPS, the relative density of 

monitoring sites is much less than that 

in the eastern U.S. This relative 

disparity in spatial density of monitors 

is exacerbated as air quality patterns in 

the mountainous west generally exhibit 

greater spatial heterogeneity due to 

dramatic elevation gradients that impact 

meteorology and air mass flow patterns. 

Similarly, water quality data coverage is 

far more comprehensive in the eastern 

U.S. relative to the west 

Measurements of NOy notably are 

lacking in both eastern and western 

acid-sensitive ecoregions. This adds 

uncertainty to the use of the AAI 

equation as the lack of NOy data limits 

efforts to evaluate air quality modeling 

of NOy that is the basis for quantifying 

factor F3 in the AAI equation. The lack 

of NOy measurements also limits efforts 

to characterize the variability and 

representativeness of modeled NOy

concentrations within and across 

ecoregions. Currently, the Agency’s 

ability to define the protection likely to 

be afforded by alternative standards (in 

terms of alternative levels and 

percentiles) is compromised by the lack 

of a full set of ambient air quality 

indicator measurements, notably 

including NOy, throughout sensitive 

ecoregions across the U.S. 

Further, obtaining measurement of the 

dominant species that comprise NOy

(HNO3, true NO2, NO, p–NO3, and PAN) 

would be useful to evaluate 

performance of NOy samplers. Beyond 

the more well known dominant 

components of NOy, research efforts 

would be needed to characterize total 

reactive nitrogen that may include 

significant amounts of organicallybound nitrogen (beyond PAN) which is 

poorly understood with regard to 

emission sources and concentration 

levels. 

Field measurements of NHx have been 

extremely limited, but have begun to be 

enhanced through the NADP’s passive 

ammonia network (AMoN). The AMoN 

measures ammonia at over 50 sites, with 

more than 35 at CASTNET locations. 

Enhanced spatial coverage of reduced 

nitrogen measurements, particularly to 

understand within and across ecoregion 

variability, and the inclusion of some 

continuous observations would provide 

a better understanding of the 

uncertainty in the F2 factor in the AAI 

equation and of the representativeness 

of modeled NHx deposition within and 

across ecoregions. 

With regard to water quality data, the 

PA notes that such data are typically 

limited relative to air quality data sets, 

and are also relatively sparse in the 

western U.S. The TIME/LTM water 

quality sampling program in the eastern 

U.S. (as described in chapter 2 of the 

PA) is an appropriate complement to 

national air monitoring programs as it 

affords consistency across water bodies 

in terms of sampling frequency and 

analysis protocols. Consideration 

should be given to extending the TIME/ 

LTM design to all acid sensitive 

ecoregions, with priority for areas in the 

western mountains that are data limited 

and showing initial signs of adversity 

particularly with respect to aquatic 

acidification. The lack of a regulatory 

requirement for TIME/LTM often 

jeopardizes funding support of this 

resource that is especially valuable and 

cost effective. While there are several 

state and local agency water quality data 

bases, it is unclear the extent to which 

differences in sampling, chemical 

analysis and reporting protocols would 

impact the use of such data for the 

purpose of better understanding the 

degree of protectiveness that would be 

afforded by an AAI-based standard 

within sensitive ecoregions across the 

country. In addition, our understanding 

of water quality in Alaska and Hawaii 

and the acid sensitivity of their 

ecoregions is particularly limited. 

Water quality data and modeling 

support the standard setting process. As 

more water bodies are sampled, the 

critical load data bases would expand, 

enabling clearer delineation of 

ecoregion representative critical loads in 

terms of the nth percentile. This would 

provide more refined characterization of 

the degree of protection afforded by a 

given standard. Longer term, the 

availability of water quality trend data 

(annual to monthly sampled) would 

support accountability assessments that 

examine if an ecoregion’s response to air 

management efforts is as predicted by 

earlier model forecasting. The most 

obvious example is the long-term 

response of water quality ANC change 

to changes in calculated AAI, 

deposition, ambient NOy and SOx

concentrations, and emissions. In 

addition, water quality trends data 

provide a basis for evaluating and 

improving the parameterizations of 

processes in critical load models 

applied at the ecoregion scale related to 

nitrogen retention and base cation 

supply. A better understanding of soil 

processes, especially in the southern 

Appalachians, would enhance efforts to 

examine the variability within 

ecoregions of the soil-based adsorption 

and exchange processes which moderate 

the supply of major cations and anions 

to surface waters and strongly influence 

the response of surface water ANC to 

changes in deposition of nitrogen and 

sulfur. 

3. Uncertainties in Modeled Processes 

As discussed in the PA, from an 

uncertainty perspective, gaps in field 

measurement data are related to 

uncertainties in modeled processes and 

in the specific application of such 

models. As noted above, processes that 

are embodied in an AAI-based standard 

are modeled using the CMAQ 

atmospheric model and steady state 

ecological models. These models are 

characterized in the ISA as being well 

established and they have undergone 

extensive peer review. Nonetheless, the 

application of these models for purposes 

of specifying the factors in the AAI 

equation, on an ecoregion scale, is a 

new application that introduces 

uncertainties, as noted below, especially 

in areas with limited observational data 

that can be used to evaluate this specific 

application. Understanding 

uncertainties in relevant modeled 

process thus involves consideration of 

the uncertainties associated with 

applying each model as well as the 

combination of these uncertainties as 

the models are applied in combination 

within the AAI framework. 

With regard to the application of 

CMAQ for purposes of use in an AAIbased standard, the modeling of dry 

deposition has been identified as having 

a relatively high degree of uncertainty. 

Due to a combination of system 

complexity and resource constraints, 

there is no routine observational basis 

for directly comparing modeled dry 

deposition and measurements. Periodic 

dry deposition flux experiments 

covering a variety of vegetation, surfaces 

and meteorology across seasons would 


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enable a more robust evaluation of 

modeled deposition of nitrogen and 

sulfur. Given the difficulty in acquiring 

dry deposition observations, it becomes 

especially important to evaluate the 

model’s ability to capture temporal and 

spatial ambient air patterns of 

individual nitrogen and sulfur species 

which are used to drive dry deposition 

calculations in models. For example, 

reducing a generally acknowledged 

positive bias in model-predicted SO2

relative to observations is especially 

relevant to the AAI-based standard, as 

SO2 deposition is a dominant 

contributor to total acidifying 

deposition in the eastern U.S. With 

respect to oxidized nitrogen, 

observations of individual NOy species 

are important as air quality models 

calculate the individual deposition of 

each species. The modeled transference 

ratios, TNOy and TSOx used in factors F3 

and F4 rely on CMAQ’s ability to 

characterize both deposition and 

concentration. Consequently, a better 

understanding of the variability of these 

factors within and across ecoregions 

could be achieved by improved 

availability of measured ambient 

concentrations and deposition 

observations. 

Steady state biogeochemical 

ecosystem modeling is used to develop 

critical load estimates that are 

incorporated in the AAI equation 

through factor F1. Consequently, the PA 

notes that an estimate of the temporal 

response of surface water ANC to 

deposition and air concentration 

changes is not directly available. 

Lacking a predicted temporal response 

impairs the ability to conduct 

accountability assessments down to the 

effects level. Accountability assessments 

would examine the response of each 

step in the emissions source through air 

concentration—deposition—surface 

water quality—biota continuum. The 

steady state assumption at the 

ecosystem level does not impair 

accountability assessments through the 

air concentration/deposition range of 

that continuum. However, in using 

steady state ecosystem modeling, 

several assumptions are made relative to 

the long-term importance of processes 

related to soil adsorption of major ions 

and ecosystem nitrogen dynamics. 

Because these models often were 

developed and applied in glaciated 

areas with relatively thin and 

organically rich soils, their applicability 

is relatively more uncertain in areas 

such as those in the non-glaciated claybased soil regions of the central 

Appalachians. Consequently, it is 

desirable to develop the information 

bases to drive simple dynamic 

ecosystem models that incorporate more 

detailed treatment of subsurface 

processes, such as adsorption and 

exchange processes and sulfate 

absorption. 

4. Applying Knowledge of Uncertainties 

An understanding of the relative 

uncertainties in a system assists in 

setting priorities for data collection 

efforts and research, with the 

expectation that such efforts would 

reduce uncertainties over time and 

afford greater confidence in applications 

of an AAI-based standard. Because of 

the uniquely wide breadth of pollutants 

and environmental media addressed by 

an AAI-based multi-pollutant standard, 

there are a wide range of uncertainties 

that are important to consider relative to 

single pollutant standards that typically 

address only direct effects of ambient air 

exposures. For an AAI-based standard, a 

reduction of the uncertainties across the 

various modeled processes at the 

ecoregion scale would lead to greater 

confidence in the degree of protection 

afforded by the standard. 

The PA notes that there is generally 

low uncertainty with regard to the 

conceptual development and related 

major components of this standard. In 

recognizing the scientific soundness of 

the basic structure of this standard, the 

PA notes that future efforts would be 

appropriately directed at expanding the 

availability of relevant data for 

ecoregion-specific evaluation and 

application of the relevant modeling of 

ecological and atmospheric processes, 

as identified above. Such efforts would 

further support consideration of an AAIbased standard and would guide field 

studies and analyses designed to 

improve the longer-term confidence in 

such a standard. 

G. CASAC Advice 

The CASAC has advised EPA 

concerning the ISA, the REA, and the 

PA. The CASAC has endorsed EPA’s 

interpretation of the science embodied 

in the ISA and the assessment 

approaches and conclusions 

incorporated in the REA. 

Most recently, CASAC has considered 

the information in the final PA in 

providing its recommendations on the 

review of the new multi-pollutant 

standard developed in that document 

and discussed above (Russell and 

Samet, 2011a). In so doing, CASAC has 

expressed general support for the 

conceptual framework of the standard 

based on the underlying scientific 

information, as well as for the 

conclusions in the PA with regard to 

indicators, form, averaging time, and 

level of the standard that are 

appropriate for consideration by the 

Agency in reaching decisions on the 

review of the secondary NAAQS for 

oxides of nitrogen and sulfur: 

The final Policy Assessment clearly sets 

out the basis for the recommended ranges for 

each of the four elements (indicator, 

averaging time, level and form) of a potential 

NAAQS that uses ambient air indicators to 

address the combined effects of oxides of 

nitrogen and oxides of sulfur on aquatic 

ecosystems, primarily streams and lakes. As 

requested in our previous letters, the Policy 

Assessment also describes the implications of 

choosing specific combinations of elements 

and provides numerous maps and tabular 

estimates of the spatial extent and degree of 

severity of NAAQS exceedances expected to 

result from possible combinations of the 

elements of the standard. 

We believe this final PA is appropriate for 

use in determining a secondary standard to 

help protect aquatic ecosystems from 

acidifying deposition of oxides of sulfur and 

nitrogen. EPA staff has done a commendable 

job developing the innovative Aquatic 

Acidification Index (AAI), which provides a 

framework for a national standard based on 

ambient concentrations that also takes into 

account regional differences in sensitivities 

of ecosystems across the country to effects of 

acidifying deposition. (Russell and Samet, 

2011a) 

The CASAC also recommended that 

as EPA moves forward in the regulatory 

process ‘‘some attention should be given 

to our residual concern that the 

available data may reflect the more 

sensitive water bodies and thus, the 

selection of percentiles of waterbodies 

to be protected could be conservatively 

biased’’ (Russell and Samet, 2011a). In 

addition, CASAC found some 

improvements could be made to the 

uncertainty analysis, as noted below. 

With respect to indicators, CASAC 

supports the use of SOx and NOy as 

ambient air indicators (discussed above 

in section III.A) and ANC as the 

ecological indicator (discussed above in 

section III.B.1): 

The use of NOy and SOx as 

atmospheric indicators of oxides of 

nitrogen and sulfur atmospheric 

concentrations is well justified. The use 

in the AAI of NOy and SOx as 

atmospheric indicators of oxides of 

nitrogen and sulfur concentrations is 

useful and corresponds with other 

efforts by EPA. As we have stated 

previously, CASAC also agrees that 

ANC is the most appropriate ecological 

indicator of aquatic ecosystem response 

and resiliency to acidification (Russell 

and Samet, 2011a). 

With respect to the form of the 

standard (discussed above in section 

III.B), CASAC stated the following: 

EPA has developed the AAI, an innovative 

‘‘form’’ of the NAAQS itself that incorporates 


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15Section 109(d)(1) requires that ‘‘* * * the 

Administrator shall complete a thorough review 

* * * and shall make such revisions in such 

criteria and standards and promulgate such new 

standards as may be appropriate under * * * 

subsection 109(b) of this section.’’ [emphasis added] 

the multi-pollutant, multi-media, 

environmentally modified, geographically 

variable nature of SOx/NOy depositionrelated aquatic acidification effects. With the 

caveats noted below, CASAC believes that 

this form of the NAAQS as described in the 

final Policy Assessment is consistent with 

and directly reflective of current scientific 

understanding of effects of acidifying 

deposition on aquatic ecosystems. (Russell 

and Samet, 2011a) 

CASAC agrees that the spatial components 

of the form in the Policy Assessment are 

reasonable and that use of Omernick’s 

ecoregions (Level III) is appropriate for a 

secondary NAAQs intended to protect the 

aquatic environment from acidification 

* * * (Russell and Samet, 2011a) 

The ‘‘caveats’’ noted by CASAC 

include a recognition of the importance 

of continuing to evaluate the 

performance of the CMAQ and 

ecological models to account for model 

uncertainties and to make the modeldependent factors in the AAI more 

transparent. In addition, CASAC noted 

that the role of DOC and its effects on 

ANC would benefit from further 

refinement and clarification (Russell 

and Samet, 2011a). While CASAC 

expressed the view that the ‘‘division of 

ecoregions into ‘sensitive’ and ‘nonsensitive’ subsets, with a more 

protective percentile applied to the 

sensitive areas, also seems reasonable’’ 

(Russell and Samet, 2011a), CASAC also 

noted that there was the need for greater 

clarity in specifying how appropriate 

screening criteria would be applied in 

assigning ecoregions to these categories. 

Further, CASAC identified potential 

biases in critical load calculations and 

in the regional representativeness of 

available water chemistry data, leading 

to the observation that a given 

percentile of the distribution of 

estimated critical loads may be 

protective of a higher percentage of 

surface waters in some regions (Russell 

and Samet, 2011a). 

With respect to averaging time 

(discussed above in section III.C), 

CASAC stated the following: 

Considering the cumulative nature of the 

long-term adverse ecological effects and the 

year-to-year variability of atmospheric 

conditions (mainly in the amount of 

precipitation), CASAC concurs with EPA that 

an averaging time of three to five years for 

the AAI parameters is appropriate. A longer 

averaging time would mask possible trends of 

AAI, while a shorter averaging time would 

make the AAI being more influenced by the 

conditions of the particular years selected. 

(Russell and Samet, 2011a) 

With respect to level as well as the 

combination of level and form as they 

are presented as alternative standards 

(discussed above in sections III.D–E), 

CASAC stated the following: 

CASAC agrees with EPA staff’s 

recommendation that the ‘‘level’’ of the 

alternative AAI standards should be within 

the range of 20 and 75 μeq/L. We also 

recognize that both the ‘‘level’’ and the form 

of any AAI standard are so closely linked in 

their effectiveness that these two elements 

should be considered together. (Russell and 

Samet, 2011a) 

When considered in isolation, it is difficult 

to evaluate the logic or implications of 

selecting from percentiles (70th to 90th) of 

the distribution of estimated critical loads for 

lakes in sensitive ecoregions to determine an 

acceptable amount of deposition for a given 

ecoregion. However, when these percentile 

ranges are combined with alternative levels 

within the staff-recommended ANC range of 

20 to 75 microequivalents per liter (μeq/L), 

the results using the AAI point to the 

ecoregions across the country that would be 

expected to require additional protection 

from acidifying deposition. Reasonable 

choices were made in developing the form. 

The number of acid sensitive regions not 

likely to meet the standard will be affected 

both by choice of ANC level and the 

percentile of the distribution of critical loads 

for lakes to meet alternative ANC levels in 

each region. These combined 

recommendations provide the Administrator 

with a broad but reasonable range of 

minimally to substantially protective options 

for the standard. (Russell and Samet, 2011a) 

CASAC also commented on EPA’s 

uncertainty analysis, and provided 

advice on areas requiring further 

clarification in the proposed rule and 

future research. The CASAC found it 

‘‘difficult to judge the adequacy of the 

uncertainty analysis performed by EPA 

because of lack of details on data inputs 

and the methodology used, and lack of 

clarity in presentation’’ (Russell and 

Samet, 2011a). In particular, CASAC 

identified the need for more thorough 

model evaluations of critical load and 

atmospheric modeling, recognizing the 

important role of models as they are 

incorporated in the form of the 

standard. In light of the innovative 

nature of the standard developed in the 

PA, CASAC identified ‘‘a number of 

areas that should be the focus of further 

research’’ (Russell and Samet, 2011a). 

While CASAC recognized that EPA staff 

was able to address some of the issues 

in the PA, they also noted areas ‘‘that 

would benefit from further study or 

consideration in potential revisions or 

modifications to the form of the 

standard.’’ Such research areas include 

‘‘sulfur retention and mobilization in 

the soils, aluminum availability, soil 

versus water acidification and 

ecosystem recovery times.’’ Further, 

CASAC encouraged future efforts to 

monitor individual ambient nitrogen 

species, which would help inform 

further CMAQ evaluations and the 

specification of model-derived elements 

in the AAI equation (Russell and Samet, 

2011a). 

H. Administrator’s Proposed 

Conclusions 

Having concluded that the existing 

NO2 and SO2 secondary standards are 

neither sufficiently protective nor 

appropriate to address depositionrelated effects associated with oxides of 

nitrogen and sulfur (section II.D above), 

the Administrator has considered 

whether it is appropriate at this time to 

set a new multi-pollutant standard for 

that purpose, with a structure that 

would better reflect the available 

science regarding acid deposition. In 

considering this, she recognizes that 

such an appropriate standard, for 

purposes of section 109(b) and (d) of the 

CAA,15 must in her judgment be 

requisite to protect public welfare, such 

that it would be neither more nor less 

stringent that necessary for that 

purpose. In particular, she has focused 

on the new standard developed in the 

PA and reviewed by CASAC, as 

discussed above. In so doing, the 

Administrator first considered the 

extent to which there is a scientific basis 

for development of such a standard, 

specifically with regard to a standard 

that would provide protection from 

deposition-related aquatic acidification 

in sensitive aquatic ecosystems in areas 

across the country. As discussed above, 

the Administrator notes that the ISA 

concludes that the available scientific 

evidence is sufficient to infer a causal 

relationship between acidifying 

deposition of nitrogen and sulfur in 

aquatic ecosystems, and that the 

deposition of oxides of nitrogen and 

sulfur both cause such acidification 

under current conditions in the U.S. 

Further, the ISA concludes that there 

are well-established water quality and 

biological indicators of aquatic 

acidification as well as well-established 

models that address deposition, water 

quality, and effects on ecosystem biota, 

and that ecosystem sensitivity to 

acidification varies across the country 

according to present and historic 

nitrogen and sulfur deposition as well 

as geologic, soil, vegetative, and 

hydrologic factors. Based on these 

considerations, the Administrator agrees 

with the conclusion in the PA, and 

supported by CASAC, that there is a 

strong scientific basis for development 


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of a standard with the general structure 

presented in the PA. 

The Administrator also recognizes 

that the conceptual framework for an 

ecologically relevant, multi-pollutant 

standard, which was initially explored 

in the REA and further developed in the 

PA, builds on the information in the 

ISA. She notes that the structure of the 

standard addresses the combined effects 

of deposition from oxides of nitrogen 

and sulfur by characterizing the linkages 

between ambient concentrations, 

deposition, and aquatic acidification, 

and that the structure of the standard 

takes into account relevant variations in 

these linkages across the country. She 

recognizes that while the standard is 

innovative and unique, the structure of 

the standard is well grounded in the 

science underlying the relationships 

between ambient concentrations of 

oxides of nitrogen and sulfur and the 

aquatic acidification related to 

deposition of nitrogen and sulfur 

associated with such ambient 

concentrations. 

While the Administrator recognizes 

the strong scientific foundation for the 

structure of an AAI-based standard, she 

also recognizes that the standard 

depends on atmospheric and ecological 

modeling, based on appropriate data, to 

specify the terms of an equation that 

incorporates the linkages between 

ambient concentrations, deposition, and 

aquatic acidification. This equation, 

which defines an aquatic acidification 

index (AAI), has the effect of translating 

spatially variable ecological effects into 

a potential national standard. With 

respect to establishing the specific terms 

of this equation, there are a number of 

inherent uncertainties and complexities 

that are relevant to the question of 

whether it is appropriate under section 

109 to set a specific AAI-based standard 

at this time, recognizing that such a 

standard must be requisite to protect 

public welfare without being either 

more or less stringent than necessary for 

this purpose. As discussed above, these 

uncertainties and complexities generally 

relate not to the structure of the 

standard, but to the quantification of the 

various elements of the standard, such 

as the F factors discussed earlier in this 

section and their representativeness at 

an ecoregion scale. These uncertainties 

and complexities currently limit efforts 

to characterize the degree of 

protectiveness that would be afforded 

by such a standard, within the ranges of 

levels and forms identified in the PA, 

and the representativeness of F factors 

in the AAI equation described above 

and in the PA. These important 

uncertainties have been generally 

categorized as limitations in available 

field data as well as uncertainties that 

are related to reliance on the application 

of ecological and atmospheric modeling 

at the ecoregion scale to specify the 

various elements of the AAI. 

With regard to data limitations, the 

Administrator observes that there are 

several important limitations in the 

available data upon which elements of 

the AAI are based. For example, while 

ambient measurements of NOy are made 

as part of a national monitoring 

network, the monitors are not located in 

locations that are representative of 

sensitive aquatic ecosystems. While air 

and water quality data are generally 

available in areas in the eastern U.S., 

there is relatively sparse coverage in 

mountainous western areas where a 

number of sensitive aquatic ecosystems 

are located. Further, even in areas where 

relevant data are available, small sample 

sizes impede efforts to characterize the 

representativeness of the available data, 

which was noted by CASAC as being of 

particular concern. Also, measurements 

of reduced forms of nitrogen are 

available from only a small number of 

monitoring sites, and emission 

inventories for reduced forms of 

nitrogen used in atmospheric modeling 

are subject to considerable uncertainty. 

With regard to uncertainties related to 

the use of ecological and atmospheric 

modeling, the Administrator notes in 

particular that model results are 

difficult to evaluate due to a lack of 

relevant observational data. For 

example, relatively large uncertainties 

are introduced by a lack of data with 

regard to pre-industrial environmental 

conditions and other parameters that are 

necessary inputs to critical load models 

that are the basis for factor F1 in the 

AAI equation. Also, observational data 

are not generally available to evaluate 

the modeled relationships between 

nitrogen and sulfur in the ambient air 

and associated deposition, which are 

the basis for the other factors (i.e., F2, 

F3, and F4) in the AAI equation. 

In combination, these limitations and 

uncertainties result in a considerable 

degree of uncertainty as to how well the 

quantified elements of the AAI standard 

would predict the actual relationship 

between varying ambient concentrations 

of oxides of nitrogen and sulfur and 

steady state ANC levels across the 

distribution of water bodies within the 

various ecoregions in the U.S. Because 

of this, there is considerable uncertainty 

as to the actual degree of protectiveness 

that such a standard would provide, 

especially for acid-sensitive ecoregions. 

The Administrator recognizes that the 

AAI equation, with factors quantified in 

the ranges discussed above and 

described more fully in the PA, 

generally performs well in identifying 

areas of the country that are sensitive to 

such acidifying deposition and 

indicates, as expected, that lower 

ambient levels of oxides of nitrogen and 

sulfur would lead to higher calculated 

AAI values. However, the uncertainties 

discussed here are critical for 

determining the actual degree of 

protection that would be afforded such 

areas by any specific target ANC level 

and percentile of water bodies that 

would be chosen in setting a new AAIbased standard, and thus for 

determining an appropriate AAI-based 

standard that meets the requirements of 

section 109. 

In considering these uncertainties, the 

Administrator notes that CASAC 

acknowledged that important 

uncertainties remain that would benefit 

from further study and data collection 

efforts, which might lead to potential 

revisions or modifications to the form of 

the standard developed in the PA. She 

also notes that CASAC encouraged the 

Agency to engage in future monitoring 

and model evaluation efforts to help 

inform the specification of modelderived elements in the AAI equation. 

Based on the above considerations, 

the Administrator has determined that it 

is not appropriate under section 109 to 

set a new multi-pollutant standard to 

address deposition-related effects of 

oxides of nitrogen and sulfur on aquatic 

acidification at this time. Setting a 

NAAQS generally involves 

consideration of the degree of 

uncertainties in the science and other 

information, such as gaps in the relevant 

data and, in this case, limitations in the 

evaluation of the application of relevant 

ecological and atmospheric models at an 

ecoregion scale. As noted above, the 

issue here is not a question of 

uncertainties about the scientific 

soundness of the structure of the AAI, 

but instead uncertainties in the 

quantification and representativeness of 

the elements of the AAI as they vary in 

ecoregions across the country. At 

present, these uncertainties prevent an 

understanding of the degree of 

protectiveness that would be afforded to 

various ecoregions across the country by 

a new standard defined in terms of a 

specific nationwide target ANC level 

and a specific percentile of water bodies 

for acid-sensitive ecoregions and thus 

prevent identification of an appropriate 

standard.. The Administrator has 

considered whether these uncertainties 

could be appropriately accounted for by 

choosing either a more or less protective 

target ANC level and percentile of water 

bodies than would otherwise be chosen 

if the uncertainties did not substantially 

limit the confidence that can 


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appropriately be ascribed to the 

quantification of the AAI elements. 

However, in the Administrator’s 

judgment, the uncertainties are of such 

nature and magnitude that there is no 

reasoned way to choose such a specific 

nationwide target ANC level or 

percentile of water bodies that would 

appropriately account for the 

uncertainties, since neither the direction 

nor the magnitude of change from the 

target level and percentile that would 

otherwise be chosen can reasonably be 

ascertained at this time. 

Based on the above considerations, 

the Administrator judges that the 

current limitations in relevant data and 

the uncertainties associated with 

specifying the elements of the AAI 

based on modeled factors are of such 

nature and degree as to prevent her from 

reaching a reasoned decision such that 

she is adequately confident as to what 

level and form (in terms of a selected 

percentile) of such a standard would 

provide any particular intended degree 

of protection of public welfare that the 

Administrator determined satisfied the 

requirements to set an appropriate 

standard under section 109. While 

acknowledging that CASAC supported 

moving forward to establish the 

standard developed in the PA, the 

Administrator also observes that CASAC 

supported conducting further field 

studies that would better inform the 

continued development or modification 

of such a standard. Given the large 

uncertainties and complexities inherent 

in quantifying the elements of such a 

standard, largely deriving from the 

unprecedented nature of the standard 

under consideration in this review, and 

having fully considered CASAC’s 

advice, the Administrator provisionally 

concludes that it is premature to set a 

new, multi-pollutant secondary 

standard for oxides of nitrogen and 

sulfur at this time, and as such she is 

proposing not to set such a new 

secondary standard. 

While it is premature to set such a 

multi-pollutant standard at this time, 

the Administrator determines that the 

Agency should undertake a field pilot 

program to gather additional data, and 

that it is appropriate that such a 

program be undertaken before, rather 

than after, reaching a decision to set 

such a standard. As described below in 

section IV, the purpose of the program 

is to collect and analyze data so as to 

enhance our understanding of the 

degree of protectiveness that would 

likely be afforded by a standard based 

on the AAI as developed in the PA. This 

will provide additional information to 

aid the Agency in considering an 

appropriate multi-pollutant standard, 

specifically with respect to the 

acidifying effects of deposition of oxides 

of nitrogen and sulfur. PA. Data 

generated by this field program will also 

support development of an appropriate 

monitoring network that would work in 

concert with such a standard to result in 

the intended degree of protection. The 

data and analyses generated as a result 

of this program will serve to inform the 

next review of the NAAQS for oxides of 

nitrogen and sulfur. The information 

generated during the field program can 

also be used to help state agencies and 

EPA better understand how an AAIbased standard would work in terms of 

the implementation of such a standard. 

Based on the above considerations, 

the Administrator is proposing not to set 

a new multi-pollutant AAI-based 

secondary standard for oxides of 

nitrogen and sulfur in this review. In 

reaching this decision, the 

Administrator recognizes that the new 

NO2 and SO2 primary 1-hour standards 

set in 2010, while not ecologically 

relevant for a secondary standard, will 

nonetheless result in reductions in 

oxides of nitrogen and sulfur that will 

directionally benefit the environment by 

reducing NOy and SOx deposition to 

sensitive ecosystems. EPA is proposing 

to revise the secondary standards by 

adding secondary standards identical to 

the NO2 and SO2 primary 1-hour 

standards set in 2010. More specifically, 

EPA is proposing a 1-hour secondary 

NO2 standard set at a level of 100 ppb 

and a 1-hour secondary SO2 standard set 

at a level of 75 ppb. While this will not 

add secondary standards of an 

ecologically relevant form to address 

deposition-related effects, it will 

directionally provide some degree of 

additional protection. This is consistent 

with the view that the current secondary 

standards are neither sufficiently 

protective nor appropriate in form, but 

that it is not appropriate to propose to 

set a new, ecologically relevant multipollutant secondary standard at this 

time, for all of the reasons discussed 

above. 

While not a basis for this decision, the 

Administrator also recognizes that a 

new, innovative AAI-based standard 

would raise significant implementation 

issues that would need to be addressed 

consistent with the CAA requirements 

for implementation-related actions 

following the setting of a new NAAQS. 

It will take time to address these issues, 

during which the Agency will be 

conducting a field pilot program to 

gather relevant data and the 

environment will benefit from 

reductions in oxides of nitrogen and 

sulfur resulting from the new NO2 and 

SO2 primary standards, as noted above, 

as well as reductions expected to be 

achieved from EPA’s Cross-State Air 

Pollution Rule and Mercury and Air 

Toxics standards. These 

implementation-related issues are 

discussed in more detail below in 

section IV.A.5. 

The Administrator solicits comment 

on all aspects of this proposed decision, 

including the framework and elements 

of a multi-pollutant standard for oxides 

of nitrogen and sulfur to address 

deposition-related effects on sensitive 

ecosystems, with a focus on aquatic 

acidification, and the uncertainties and 

complexities associated with the 

development of such a standard at this 

time. The Administrator also solicits 

comment on the field pilot program and 

related monitoring methods as 

discussed below in section IV. 

IV. Field Pilot Program and Ambient 

Monitoring 

This section describes EPA’s plans for 

a field pilot program and the evaluation 

of monitoring methods for ambient air 

indicators of NOy and SOx to implement 

the Administrator’s decision to 

undertake such a field monitoring 

program in conjunction with her 

decision to propose not to set a new 

multi-pollutant secondary standard in 

this review, as discussed above in 

section III.H. As noted above and 

discussed below in section IV.A, the 

field pilot program is intended to collect 

and analyze data so as to enhance our 

understanding of the degree of 

protectiveness that would likely be 

afforded by a standard based on the AAI 

as developed in the PA. Data generated 

by this field program would also 

support development of an appropriate 

monitoring network that would work in 

concert with such a standard to result in 

the intended degree of protection. As 

discussed below in section IV.B, the 

evaluation of monitoring methods 

focuses on the development of Federal 

Reference Methods/Federal Equivalent 

Methods (FRM/FEM) for NOy and SOx. 

The EPA notes that the monitoring 

program described here is intended to 

be coordinated with EPA’s CASTNET as 

a supplement to existing monitoring 

programs and is beyond the scope of the 

current CASTNET program. 

A. Field Pilot Program 

This section presents the objectives of 

a field pilot program (section IV.A.1) 

that would gather relevant field data 

over a 5-year period in a sample of three 

to five sensitive ecoregions across the 

country. An overview of the scope and 

structure of the field program, with a 

focus on measurements of ambient air 

indicators of oxides of nitrogen and 


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sulfur, is presented in section IV.A.2. 

Section IV.A.3 explains the role of 

additional complementary 

measurements beyond the ambient air 

indicators that would be included in the 

program, and section IV.A.4 discusses a 

parallel longer-term research agenda, 

both of which are guided by the 

uncertainties discussed above in section 

III. Section IV.A.5 identifies 

implementation challenges presented by 

an AAI-based standard that could be 

addressed in parallel with a field pilot 

program. Section IV.A.6 discusses 

engagement with stakeholder groups as 

part of the planned pilot program. 

1. Objectives 

Consideration of a new multipollutant standard to address 

deposition-related effects on sensitive 

aquatic ecoregions raises unique 

challenges relative to those typically 

raised in reviews of existing NAAQS for 

which an established network of FRM/ 

FEM monitors, designed to measure the 

indicator pollutant, is generally 

available. The primary goal of this field 

pilot program, and the related 

monitoring program discussed in 

section IV.B, is to enhance our 

understanding of the degree of 

protectiveness that would likely be 

afforded by a standard based on the 

AAI, as described above in section III, 

so as to aid the Agency in considering 

an appropriate multi-pollutant standard 

that would be requisite to protect public 

welfare consistent with section 109 of 

the CAA, through the following 

objectives: 

(1) Evaluate measurement methods for 

the ambient air indicators of NOy and 

SOx and consider designation of such 

methods as FRMs; 

(2) Examine the variability and 

improve characterization of 

concentration and deposition patterns of 

NOy and SOx, as well as reduced forms 

of nitrogen, within and across a number 

of sensitive ecoregions across the 

country; 

(3) Develop updated ecoregionspecific factors (i.e., F1 through F4) for 

the AAI equation based in part on new 

observed air quality data within the 

sample ecoregions as well as on updated 

nationwide air quality model results 

and expanded critical load data bases, 

and explore alternative approaches for 

developing such representative factors; 

(4) Calculate ecoregion-specific AAI 

values using observed NOy and SOx data 

and updated ecoregion-specific factors 

to examine the extent to which the 

sample ecoregions would meet a set of 

alternative AAI-based standards; 

(5) Develop air monitoring network 

design criteria for an AAI-based 

standard; 

(6) assess the use of total nitrate 

measurements as a potential alternative 

indicator for NOy; 

(7) Support related longer-term 

research efforts, including 

enhancements to and evaluation of 

modeled dry deposition algorithms; and 

(8) Facilitate stakeholder engagement 

in addressing implementation issues 

associated with possible future adoption 

of an AAI-based standard. 

2. Overview of Field Pilot Program 

The CASTNET program (Figure IV–1) 

affords an available infrastructure 

relevant to an AAI-based standard, 

given the location of sites in some acidsensitive ecoregions and various 

measurements of sulfur and nitrogen 

species. The EPA plans to use 

CASTNET sites in selected acidsensitive ecoregions to serve as the 

platform for this pilot program, 

potentially starting in late 2012 and 

extending through 2018. The CASTNET 

sites in three to five ecoregions in acidsensitive areas would collect NOy and 

SOx (i.e., SO2 and p-SO4) measurements 

over a 5-year period. The initial step in 

developing a data base of observed 

ambient air indicators for oxides of 

nitrogen and sulfur requires the 

addition of NOy samplers at the pilot 

study sites so that a full complement of 

indicator measurements are available to 

calculate AAI values. These CASTNET 

sites would also be used to make 

supplemental observations useful for 

evaluation of CMAQ’s characterization 

of factors F2 –F4 in the AAI equation. 

The selected ecoregions would 

account for geographic variability by 

including regions from across the U.S., 

including the east, upper midwest and 

west. Each selected region would have 

at least two existing CASTNET sites. 

Each of the pilot CASTNET sites would 

be used to evaluate the performance of 

the established methods, data retrieval 

and reporting procedures used in the 

AAI equation. 


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Over the course of this 5-year pilot 

program, the most current national air 

quality modeling, based on the most 

current national emissions inventory, 

would be used to develop an updated 

set of F2—F4 factors. A parallel multiagency national critical load data base 

development effort would be used as the 

basis for calculating updated F1 factors. 

As discussed above in section III.B, 

these factors would be based on average 

parameter values across an ecoregion. 

Using this new set of F factors, 

observations of NOy and SOx derived 

from the pilot program, averaged across 

each ecoregion, would be used to 

calculate AAI values in the sample 

ecoregions. The data from the pilot 

program would also be used to examine 

alternative approaches to generating 

representative air quality values, such as 

examining the appropriateness of spatial 

averaging in areas of high spatial 

variability. 

3. Complementary Measurements 

Complementary measurements may 

be performed at some sites in the pilot 

network to reduce uncertainties in the 

recommended methods and better 

characterize model performance and 

application to the AAI. The CASAC Air 

Monitoring and Methods Subcommittee 

(AMMS) advised EPA that such 

supplemental measurements were of 

critical importance in a field 

measurement program related to an 

AAI-based standard (Russell and Samet, 

2011b). 

Candidate complementary 

measurements to address sulfur, in 

addition to those provided by the 

CASTNET filter pack (CFP), include 

trace gas continuous SO2 and speciated 

PM2.5 measurements. The co-located 

deployment of a continuous SO2

analyzer with the CFP for SO2 will 

provide test data for determining 

suitability of continuous SO2

measurements as a Federal Equivalent 

Method (FEM), as well as producing 

valuable time series data for model 

evaluation purposes. The weekly 

averaging time provided by the CFP 

adequately addresses the annual-average 

basis of an AAI-based secondary 

standard, but would not be applicable to 

short-term (i.e., 1-hour) averages 

associated with the primary SO2

standard. Conversely, because of the 

low concentrations associated with 

many acid-sensitive ecoregions, existing 

SO2 Federal Reference Methods (FRMs) 

designated for use in determining 

compliance with the primary standard 

would not necessarily be appropriate for 

use in conjunction with an AAI-based 

secondary standard. 

Co-locating the PM2.5 sampler used in 

the EPA Chemical Speciation Network 

and the Interagency Monitoring of 

Protected Visual Environments 

(IMPROVE) network at pilot network 

sites would allow for characterizing the 

relationship between the CFP-derived pSO4 and the speciation samplers used 

throughout the state and local air 

quality networks. Note that CASTNET 

already has several co-located IMPROVE 

chemical speciation samplers. Because 

the AAI equation is based on 

concentration of p-SO4, the original 

motivation for capturing all particle size 

fractions is not as important relative to 

simply capturing the concentration of 

total p-SO4. 

Candidate measurements to 

complement oxidized nitrogen 

measurements, in addition to the CFP, 

include a mix of continuous and 

periodic sampling for the dominant NOy

species, namely NO, true NO2, PAN, 


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HNO3, and p-NO3. While there are 

several approaches to acquiring these 

measurements, perhaps the most 

efficient strategy would take advantage 

of the available CFP for total nitrate, and 

add a three-channel chemiluminescence 

instrument that will cycle between NOy, 

true NO2 and NO by adding photolytic 

detection for true NO2. Other options for 

measuring true NO2 would include 

adding either a stand-alone photolytic 

or cavity ring-down spectroscopy 

instrument. Measurements of PAN may 

be acquired either on a periodic basis 

through canister sampling and 

subsequent laboratory analysis or 

through emerging in-situ sampling and 

analysis methods. Although the CFP 

yields a reliable measurement of total 

nitrate, the t-NO3 (i.e., the sum of HNO3

and p-NH4) value, strong consideration 

may be given to direct measurement of 

HNO3, which has the highest deposition 

velocity of all the dominant NOy

species. Similar to the use of continuous 

SO2 data, these speciated NOy data serve 

two purposes: evaluating total NOy

instrument behavior and evaluating air 

quality models. The measurement of 

individual NOy species can be used to 

generate site-specific NOy values for 

comparison to modeled NOy, and will 

likely provide insight into and 

improvement of modeled dry 

deposition. 

The CASAC AMMS (Russell and 

Samet, 2011b) recommended that EPA 

consider the use of t-NO3 obtained from 

CASTNET sampling as an indicator for 

NOy, reasoning that t-NO3 is typically a 

significant fraction of deposited 

oxidized nitrogen in rural environments 

and CASTNET measurements are 

widely available. Collection of this data 

would support further consideration of 

using the CFP for t-NO3 as the indicator 

of oxides of nitrogen for use in an AAIbased secondary standard. 

The CASAC AMMS also 

recommended that total NHx (NH3 and 

p-NH4) be considered as a proxy for 

reduced nitrogen species, reasoning that 

the subsequent partitioning to NH3 and 

p-NH4 may be estimated using 

equilibrium chemistry calculations. 

Reduced nitrogen measurements are 

used to evaluate air quality modeling 

which is used in generating factor F2. 

Additional studies are needed to 

determine the applicability of NHx

measurements and calculated values of 

NH3 and NH4 to the AAI. 

The additional supplemental 

measurements of speciated NOy, 

continuous SO2 and NHx will be used in 

future air quality modeling evaluation 

efforts. Because there often is significant 

lag in the availability of contemporary 

emissions data to drive air quality 

modeling, the complete use of these 

data sets will extend beyond the 5-year 

collection period of the pilot program. 

Consequently, the immediate 

application of those data will address 

instrument performance comparisons 

that explore the feasibility of using 

continuous SO2 instruments in rural 

environments, and using the speciated 

NOy data to assess NOy instrument 

performance. Although contemporary 

air quality modeling will lag behind 

measurement data availability, the 

observations can be used in deposition 

models to compare observed 

transference ratios with the previously 

calculated transference ratios to test 

temporal stability of the ratios. 

An extended water quality sampling 

effort should parallel the air quality 

measurement program to address some 

of the uncertainties related to factor F1 

and the representativeness of the nth 

percentile critical load as discussed in 

section III.B.5.b.i. The objective of the 

water quality sampling would be to 

develop a larger data base of critical 

loads in each of the pilot ecoregions 

such that the nth percentile can 

adequately be characterized in terms of 

representing all water bodies. 

Opportunities to leverage and perhaps 

enhance existing ecosystem modeling 

efforts enabling more advanced critical 

load modeling and improved methods 

to estimate base cation production 

would be pursued. For example, areas 

with ongoing research studies 

producing data for dynamic critical load 

modeling would be considered when 

selecting the pilot ecoregions. 

4. Complementary areas of research 

The EPA recognizes that a source of 

uncertainty in an AAI-based secondary 

standard that would not be directly 

addressed in the pilot program stems 

from the uncertainty in the model used 

to link atmospheric concentrations to 

dry deposition fluxes. Currently, there 

are no ongoing direct dry deposition 

measurement studies at CASTNET sites 

that can be used to evaluate modeled 

results. It was strongly recommended by 

CASAC AMMS that a comprehensive 

sampling-intensive study be conducted 

in at least one, preferably two sites in 

different ecoregions to assess 

characterization of dry deposition of 

sulfur and nitrogen. These sites would 

be the same as those for the 

complementary measurements 

described above, but they would afford 

an opportunity to also complement dry 

deposition process research that benefits 

from the ambient air measurements 

collected in the pilot program. The 

concerns regarding uncertainties 

underlying an AAI-based secondary 

standard suggest that research that 

includes dry deposition measurements 

and evaluation of dry deposition models 

should be a high priority. 

Similar leveraging should be pursued 

with respect to ecosystem research 

activities. For example, studies that 

capture a suite of soil, vegetation, 

hydrological, and water quality 

properties that can help evaluate more 

advanced critical load models would 

complement the atmospheric-based 

pilot program. In concept, such studies 

could provide the infrastructure for true 

multi-pollutant, multi-media ‘‘super’’ 

sites assuming the planning, 

coordination, and resource facets can be 

aligned. While this discussion 

emphasizes the opportunity of 

leveraging ongoing research efforts, 

consideration could be given to 

explicitly including related research 

components directly in the pilot 

program. 

5. Implementation challenges 

The CAA requires that once a NAAQS 

is established, designation and 

implementation must move forward. 

With a standard as innovative as the 

AAI-based standard considered in this 

review, the Administrator believes that 

its success will be greatly improved if, 

while additional data are being 

collected to reduce the uncertainties 

discussed above, the implementing 

agencies and other stakeholders have an 

opportunity to discuss and thoroughly 

understand how such a standard would 

work. And since, as noted above, 

emissions reductions that are 

directionally correct to reduce aquatic 

acidification will be occurring as a 

result of other CAA programs, the 

Administrator believes that this period 

of further discussion will not delay 

progress but will ensure that once 

implementation is triggered, agencies 

will be prepared to implement it 

successfully. 

Consideration of an AAI-based 

secondary standard for oxides of 

nitrogen and sulfur would present 

significant implementation challenges 

because it involves multiple, regionallydispersed pollutants and relatively 

complex compliance determinations 

based on regionally variable levels of 

NOy and SOx concentrations that would 

be necessary to achieve a national ANC 

target. The anticipated implementation 

challenges fall into three main 

categories: monitoring and compliance 

determinations for area designations, 

pre-construction permit application 

analyses of individual source impacts, 

and State Implementation Plan (SIP) 

development. Several overarching 

implementation questions that we 


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anticipate will be addressed in parallel 

with the field pilot program’s five-year 

data collection period include: 

(1) What are the appropriate 

monitoring network density and siting 

requirements to support a compliance 

system based on ecoregions? 

(2) Given the unique spatial nature of 

the secondary standard (e.g., 

ecoregions), what are the appropriate 

parameters for establishing 

nonattainment areas? 

(3) How can new or modified major 

sources of oxides of nitrogen and oxides 

of sulfur emissions assess their ambient 

impacts on the standard and 

demonstrate that they are not causing or 

contributing to a violation of the 

NAAQS for preconstruction permitting? 

To what extent does the fact that a 

single source may be impacting multiple 

areas, with different acid sensitivities 

and variable levels of NOy and SOx

concentrations that would be necessary 

to achieve a national ANC target, 

complicate this assessment and how can 

these additional complexities best be 

addressed? 

(4) What additional tools, 

information, and planning structures are 

needed to assist states with SIP 

development, including the assessment 

of interstate pollutant transport and 

deposition? 

(5) Would transportation conformity 

apply in nonattainment and 

maintenance areas for this secondary 

standard, and, if it does, would 

satisfying requirements that apply for 

related primary standards (e.g., ozone, 

PM2.5, and NO2) be demonstrated to 

satisfy requirements for this secondary 

standard? 

6. Final Monitoring Plan Development 

and Stakeholder Participation 

The existing CASTNET sampling site 

infrastructure provides an effective 

means of quickly and efficiently 

deploying a monitoring program to 

support potential implementation of an 

AAI-based secondary standard, and also 

provides an additional opportunity for 

federally managed networks to 

collaborate and support the states, local 

agencies and tribes (SLT) in determining 

compliance with a secondary standard. 

A collaborative effort would help to 

optimize limited federal and SLT 

monitoring funds and would be 

beneficial to all involved. The 

CASTNET is already a stakeholderbased program with over 20 participants 

and contributors, including federal, 

state and tribal partners. 

The CASAC AMMS generally 

endorsed the technical approaches used 

in CASTNET, but concerns were raised 

by individual representatives of state 

agencies concerning the perception of 

EPA-controlled management aspects of 

CASTNET and data ownership. 

Potential approaches to resolve these 

issues will be developed and evaluated 

in existing National Association of 

Clean Air Agencies (NACAA)/EPA 

ambient air monitoring workgroups. The 

EPA Office of Air and Radiation (which 

includes the Office of Air Quality 

Planning Standards, OAQPS; and the 

Office of Atmospheric Program’s Clean 

Air Markets Division, OAP–CAMD), and 

their partners on the NACAA monitor 

steering committee will develop a 

prioritized specific plan that identifies 

the three to five ecoregions and the 

instrumentation to be deployed. The 

EPA anticipates that a cost estimate of 

the plan with priorities and options will 

be developed by January, 2012. 

Although this pilot program is focused 

on data collection, the plan will include 

details of the data analysis approaches 

as well as a vehicle that incorporates 

engagement from those within EPA and 

SLTs to foster progress on the 

implementation questions noted above 

in section IV.A.5. 

If an AAI-based secondary standard 

were to be set in the future, deployment 

of a full national network would follow 

the pilot monitoring program. The 

number of sites deployed in the network 

will lead to increased confidence in 

capturing spatial patterns of air quality. 

Recognizing that this section presents 

the general elements of the field pilot 

programs, EPA intends to develop a 

more detailed field pilot program plan 

through a process that will engage the 

air quality management and research 

(atmospheric and ecosystem) 

communities, as well as other federal 

agencies, state and local agencies, and 

non-government based centers of 

expertise. The EPA is seeking comment 

and input on all aspects of this field 

pilot program. 

B. Evaluation of Monitoring Methods 

The EPA generally relies on 

monitoring methods that have been 

designated as FRMs or FEMs for the 

purpose of determining the attainment 

status of areas with regard to existing 

NAAQS. Such FRMs or FEMs are 

generally required to measure the air 

quality indicators that are compared to 

the level of a standard to assess 

compliance with a NAAQS. Prior to 

their designation by EPA as FRM/FEMs 

through a rulemaking process, these 

methods must be determined to be 

applicable for routine field use and need 

to have been experimentally validated 

by meeting or exceeding specific 

accuracy, reproducibility, and reliability 

criteria established by EPA for this 

purpose. As discussed above in section 

III.A, the ambient air indicators being 

considered for use in an AAI-based 

standard include SO2, particulate sulfate 

(p-SO4), and total reactive oxides of 

nitrogen (NOy). 

The CASTNET provides a well 

established infrastructure that would 

meet the basic location and 

measurement requirements of an AAIbased secondary standard given the 

rural placement of sites in acid sensitive 

areas. In addition, CFPs currently 

provide very economical weekly, 

integrated average concentration 

measurements of SO2, p-SO4, 

ammonium ion (NH4) and t-NO3, the 

sum of HNO3 and p-NO3. 

While routinely operated instruments 

that measure SO2, p-SO4, NOy and/or tNO3 exist, instruments that measure pSO4, NOy, t-NO3, or the CFP for SO2

have not been designated by EPA as 

FRMs or FEMs. The EPA’s Office of 

Research and Development has initiated 

work that will support future FRM 

designations by EPA for SO2 and p-SO4

measurements based on the CFP. Such 

a designation by EPA could be done for 

the purpose of facilitating consistent 

research related to an AAI-based 

standard and/or in conjunction with 

setting and supporting an AAI-based 

secondary standard. 

Based on extensive review of 

literature and available data, the EPA 

has identified potential methods that 

appear suitable for measuring each of 

the three components of the indicators. 

These three methods are being 

considered as new FRMs to be used for 

measuring the ambient concentrations 

of the three components that would be 

needed to determine compliance with 

an AAI-based secondary standard. 

For the SO2 and p-SO4 measurements, 

EPA is considering the CFP method, 

which provides weekly average 

concentration measurements for SO2

and p-SO4. This method has been used 

in the EPA’s CASTNET monitoring 

network for 15 years, and strongly 

indicates that it will meet the 

requirements for use as an FRM for the 

SO2 and p-SO4 concentrations for an 

AAI-based secondary standard. 

Although the CFP method would 

provide measurements of both the SO2

and p-SO4 components in a unified 

sampling and analysis procedure, 

individual FRMs will be considered for 

each. The EPA recognizes that an 

existing FRM to measure SO2

concentrations using ultra-violet 

fluorescence (UVF) exists (40 CFR Part 

50, Appendix A–1) for the purpose of 

monitoring compliance for the primary 

SO2 NAAQS. However, several factors 

suggest that the CFP method would be 


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superior to that UVF FRM for 

monitoring compliance with an AAIbased secondary standard and will be 

discussed in more detail below. 

For monitoring the NOy component, a 

continuous analyzer for measuring NOy

is commercially available and is 

considered to be suitable for use as an 

FRM. This method is similar in design 

to the existing NO2 FRM (described in 

40 CFR Part 50, Appendix F), which is 

based on the ozone chemiluminescence 

measurement technique. The method is 

adapted to and further optimized to 

measure all NOy. However, this NOy

method requires further evaluation 

before it can be fully confirmed as a 

suitable FRM. The EPA is currently 

completing a full scientific assessment 

of the NOy method to determine 

whether it would be appropriate to 

consider for designation by EPA as an 

FRM. Specific details on these three 

methods are given below. 

On February 16, 2011, EPA presented 

this set of potential FRMs to the CASAC 

AMMS for their consideration and 

comment. In response, the CASAC 

AMMS stated that, overall, it believes 

that EPA’s planned evaluation of 

methods for measuring NOy, SO2 and pSO4 as ambient air indicators is a 

suitable approach in concept. On 

supporting the CFP method as a 

potential FRM for SO2, CASAC stated 

that they felt that the CFP is adequate 

for measuring long-term average SO2 gas 

concentrations in rural areas with low 

levels (less than 5 parts per billion by 

volume (ppbv)) and is therefore suitable 

for consideration as an FRM. For p-SO4, 

CASAC generally supports the use of 

the CFP as a potential FRM for 

measuring p-SO4 for an AAI-based 

secondary standard. The method has 

been relatively well-characterized and 

evaluated, and it has a documented, 

long-term track record of successful use 

in a field network designed to assess 

spatial patterns and long-term trends. 

On supporting the photometric NOy

method as a potential FRM, CASAC 

concluded that the existing NOy method 

is generally an appropriate approach for 

the indicator. However, CASAC agrees 

that additional characterization and 

research is needed to fully understand 

the method in order to designate it as a 

FRM. The EPA is now soliciting public 

comment on these methods as to their 

adequacy, suitability, and relative 

merits as FRMs for purposes of 

monitoring to determine compliance 

with an AAI-based secondary standard. 

1. Potential FRMs for SO2 and p-SO4

The CFP is a combined, integrated 

sampling and analysis method based on 

the well-established measurement 

technology that has been used 

extensively in EPA’s CASTNET 

monitoring network (see http:// 

www.epa.gov/castnet). This method is 

in current use at over 80 monitoring 

sites and has been in use at not less than 

40 sites for over 15 years. This method 

employs a relatively simple and 

inexpensive sampler and uses four 47- 

mm filters placed in an open-faced filter 

pack to simultaneously collect 

integrated filter samples for the SO2 and 

p-SO4 components. In addition, the CFP 

is also capable of the collection of tNO3, the sum of HNO3 and p-NO3. 

The first stage of the filter pack 

assembly contains a Teflon® filter that 

collects p-SO4

2¥ and p-NO3, the second 

stage contains a nylon filter that collects 

SO2 (as SO4

2¥) and HNO3, and the third 

stage contains two cellulose fiber filters 

impregnated with potassium carbonate 

(K2CO3) that collect any remaining SO2

(as SO4

2¥). The sampler collects 1-week 

integrated samples at a very low, 

controlled flow rate (1.5 or 3 L/min) in 

an attempt simulate actual deposition. 

Weekly averaged SO2 and p-SO4

concentrations could then be averaged 

over a 1-year period to calculate annual 

average values. 

Upon sample completion, the speciesspecific filters are extracted, with 

subsequent analysis by the wellestablished and documented ion 

chromatographic (IC) analytical 

technique. During the IC analysis, an 

aliquot of a filter extract is injected into 

a stream of eluent (ion chromatography 

mobile phase, generally a millimolarstrength solution of carbonatebicarbonate) and passed through a series 

of ion exchangers. The anions of interest 

are separated on the basis of their 

relative affinities for a low capacity and 

the strongly basic anion exchanger 

(guard and separator column). The 

separated anions are directed onto a 

cation exchanger (suppressor column) 

where they are converted to their highly 

conductive acid form, and the eluent is 

converted to a weakly conductive form. 

The now-separated anions, each in their 

acid form, are measured by 

conductivity. They are identified on the 

basis of retention time compared to that 

of standards and quantified by 

measurement of peak area compared to 

the peak areas of calibration standards. 

Calibration and quality assurance for 

the method are applied to the sample 

filters, the analytical processes, and the 

flow rate measurement and control 

aspects of the sampler. Overall method 

performance is typically assessed with 

collocated samplers. These quality 

assurance techniques are routinely used 

and have proved adequate for other 

types of FRMs and equivalent methods 

in air monitoring network service. 

The measurement and analytical 

procedures and past performance data 

associated with the CFP method are 

well documented and available through 

Quality Assurance Performance Plans 

(QAPPs), Standard Operating 

Procedures (SOPs) and annual reports 

(US EPA, 2010a and 2010b). The 

accumulated database on the CFP 

method is substantial and indicates that 

the method is sound, stable and has 

good reliability in routine, field 

operation. Data quality assessment 

results show the method to have good 

reproducibility, with collocated and 

analytical precision values in the range 

of 2 percent to 10 percent (excluding 

very low concentration measurements 

near the method detection limits; US 

EPA 2010b). 

Data quality objectives (DQOs) for a 

new FRM would be based upon current 

DQOs being used for this method by 

EPA’s OAP/CAMD and the NPS, the 

federal managers of CASTNET (US EPA, 

2010a). In its current state, the CFP 

method is expected to meet or exceed 

(as past CASTNET data have indicated; 

US EPA, 2010b) the expected FRM 

DQOs, even when deployed in new 

monitoring networks outside of 

CASTNET. In addition, CASTNET 

samples have agreed favorably with 

other measures of SO2 and p-SO4 in 

comparison studies. For example, in 

direct comparison with an annular 

denuder sampler (ADS) method, 

CASTNET/ADS ratios for SO2 and pSO4 were generally on the order of 0.9– 

1.1 (Lavery et al, 2009; Sickles et al, 

1999; Sickles et al, 2008), thus 

illustrating the accuracy of the CFP 

method in the determination of longterm average SO2 and p-SO4

concentrations. The EPA believes that 

the CFP method would be fully 

adequate as an FRM in determining 

yearly average SO2 and p-SO4

concentrations for compliance 

determination purposes. 

The EPA recognizes that an existing 

FRM for SO2 has proven adequate for 

the purposes of monitoring compliance 

for the primary SO2 NAAQS, 

specifically the newly-promulgated 1- 

hour standard. However, this FRM is 

better suited to the shorter-term, higher 

concentration primary and secondary 

SO2 NAAQS, and there is substantial 

uncertainty as to the adequacy of this 

SO2 FRM for monitoring the lower 

concentrations relevant to determining 

compliance with an AAI-based 

secondary standard. The performance 

specifications for SO2 FRM analyzers 

(40 CFR Part 53, Table B–1) require a 

lower detectable limit (LDL) of 0.002 


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ppm for the standard measurement 

range and 0.001 ppm for the lower 

measurement range. These requirements 

correspond to mass per unit volume 

concentrations of 5.24 and 2.62 μg/m3, 

respectively. Analysis of 2009 

CASTNET data shows that of the 84 

CASTNET sampling sites, 63 measured 

annual average SO2 concentrations 

below even the lower of these LDL 

requirements of 2.62 μg/m3 for the lower 

range SO2 FRM (US EPA, 2010a). In 

addition, 11 of the 84 sites measured 

annual (2009) average SO2

concentrations very near or below the 

manufacturers’ reported detection limits 

for trace level UVF SO2 monitors. 

Further, it is likely that the number of 

sites with annual average SO2

concentration below both the SO2 FRM 

LDL and the manufacturers reported 

detection limits will increase due to 

expected declines in mean SO2

concentrations (US EPA, 2010b). For 

these reasons, EPA is considering the 

CFP method for use as the FRM for 

monitoring the SO2 component of an 

ambient air indicator for oxides of 

sulfur, with a recommendation for 

additional study and data collection to 

evaluate further the possible 

applicability of the continuous UVF SO2

FRM for this purpose. 

2. Potential FRM for NOy

Atmospheric concentrations of NOy

are measured continuously by an 

analyzer that photometrically measures 

the light intensity, at wavelengths 

greater than 600 nanometers (nm), 

resulting from the chemiluminescent 

reaction of ozone (O3) with NO in 

sampled air. This method is very similar 

to the chemiluminescence NO/NO2

analyzers widely used to collect NO2

monitoring data for determining 

compliance with the NO2 NAAQS. The 

various oxides of nitrogen species, 

excluding NO, are first quantitatively 

reduced to NO by means of a catalytic 

converter. These species include NO2, 

HNO2, PANs, HNO3 and p-NO3. The 

NO, which commonly exists in ambient 

air, passes through the converter 

unchanged, and, when combined with 

the NO resulting from the catalytic 

conversion of the other oxides of 

nitrogen, a measurement of the total 

NOy concentration results. To maximize 

the conversion of the more chemically 

active oxides of nitrogen species, the 

converter is located externally, at or 

near the air sample inlet probe. This 

location minimizes losses of these active 

species that could otherwise occur from 

chemical reactions and wall losses in 

the sample inlet line. 

The NOy analyzer is a suitable, 

commercially produced continuous 

chemiluminescence analyzer that 

includes an ozone generator, a reaction 

cell, a photometric detector, wavelength 

filters as necessary to reduce sensitivity 

to wavelengths below 600 nanometer 

(nm), a pump and flow control system 

to draw atmospheric air through the 

converter and into the reaction cell, a 

suitable converter, a system to control 

the operation of the analyzer, and 

appropriate electronics to process and 

quantitatively scale the photometric 

signals. The converter contains a 

catalyst such as molybdenum and is 

heated to an optimum temperature 

designed to optimize the conversion of 

the various oxides of nitrogen to NO. It 

is connected to the analyzer via suitable 

lengths of Teflon® tubing. Hourly NOy

measurements obtained by the analyzer 

would be averaged over the same 7-day 

period used by the CFP method to 

measure the SO2 and p-SO4

components, with further averaging over 

a 1-year period. 

Commercial NOy analyzers are 

currently available, and the analyzers 

have been used for a variety of 

monitoring applications. During the 

2006 TexAQS Radical and Aerosol 

Measurement Project (TRAMP), Luke et 

al., 2010, compared measured NOy

concentrations obtained with an NOy

instrument based upon the above 

mentioned methodology with the sum 

of measured individual NOy species 

(i.e., NOyi = 

NO+NO2+HNO3+PANs+HNO2+p-NO3). 

This comparison yielded excellent 

overall agreement during both day 

([NOy](ppb) = [NOyi](ppb) × 1.03¥0.42; 

r2 = 0.9933) and night time ([NOy](ppb) 

= [NOyi](ppb) × 1.01¥0.18; r2 = 0.9975) 

periods (Luke et al, 2010). The results of 

this study show that this NOy method is 

capable of the accurate determination of 

all the atmospherically relevant NOy

components, resulting in an accurate 

determination of total NOy

concentrations. The NOy instruments 

have been routinely operated in 

networks such as SouthEastern Aerosol 

Research and Characterization 

(SEARCH), dating back several years. In 

addition, state monitoring agencies 

across the U.S. have begun, starting in 

2009, the routine operation of 

commercially available NOy

instrumentation in anticipation of EPA’s 

NCore network transitioning to full 

operation in 2011. 

These initial assessments described 

above are promising and indicate that 

the photometric NOy method appears to 

be accurate, reliable, and capable of 

routine network operation. As a result, 

the method is most likely capable for 

use as an FRM for determining 

atmospheric NOy concentrations as a 

component in determining compliance 

with an AAI-based secondary standard. 

However, as described below, this 

continuous method for NOy requires 

additional time for further evaluation 

before it can be fully confirmed for 

adoption as a FRM. The EPA has 

identified measurement uncertainties 

and some remaining science questions 

associated with this method. Among 

these are: (a) The ability of the method 

to capture all components of NOy

relevant to nitrogen deposition, (b) the 

efficiency of the molybdenum converter 

in converting all oxides of nitrogen to 

NO for detection (excluding NO2, as this 

conversion is already well documented), 

(c) appropriate inlet height 

specifications to minimize any bias 

associated with vertical concentration 

gradients of key NOy components, (d) 

identification and quantification of 

potential measurement interferences in 

the NOy determination, and (e) 

development and demonstration of 

effective calibration/challenge 

procedures to best represent the various 

mixtures of NOy components that are 

expected to be present in the different 

air sheds across the U.S. 

To address these NOy method 

uncertainties and to fully assess this 

method for use as the NOy FRM, EPA 

has developed a detailed research plan 

(Russell and Samet, 2011b) which was 

presented to the CASAC AMMS on 

February 16, 2011. In response, CASAC 

recognized the need for, and supported 

the general outline of EPA’s research 

plan to evaluate the NOy method for 

potential designation as an FRM (US 

EPA, 2011). In addition, the CASAC 

AMMS suggested additional areas of 

research associated with the 

photometric NOy method that warrant 

further assessment prior to final 

designation of the method as the NOy

FRM. These include operation of the 

method during extremely low 

temperature conditions to investigate 

possible condensation in sample lines, 

method detection limits relative to low 

levels expected in remote areas, and 

ambient-based method evaluations in 

various air sheds across the U.S. In 

response to these CASAC AMMS 

suggestions, EPA is carrying out studies, 

in addition to the tasks outlined in the 

research plan, for the NOy method. The 

results of these studies will likely take 

a year or more to become available. As 

noted previously, EPA anticipates that 

these results will be favorable and will 

confirm the adequacy of the NOy

method as a suitable FRM for 

determining compliance with an AAIbased secondary standard. 


jlentini on DSK4TPTVN1PROD with PROPOSALS3
46142 Federal Register / Vol. 76, No. 147 / Monday, August 1, 2011 / Proposed Rules 

V. Statutory and Executive Order 

Reviews 

A. Executive Order 12866: Regulatory 

Planning and Review and Review and 

Executive Order 13563: Improving 

Regulation and Regulatory Review 

Under Executive Order 12866 (58 FR 

51735, October 4, 1993), this action is a 

‘‘significant regulatory action.’’ 

Accordingly, EPA submitted this action 

to the Office of Management and Budget 

(OMB) for review under Executive 

Orders 12866 and 13563 (76 FR 3821, 

January 21, 2011), and any changes 

made in response to OMB 

recommendations have been 

documented in the docket for this 

action. 

B. Paperwork Reduction Act 

This action does not impose an 

information collection burden under the 

provisions of the Paperwork Reduction 

Act, 44 U.S.C. 3501 et seq. Burden is 

defined at 5 CFR 1320.3(b). There are no 

information collection requirements 

directly associated with the 

establishment of a NAAQS under 

section 109 of the CAA. 

C. Regulatory Flexibility Act 

For purposes of assessing the impacts 

of today’s rule on small entities, small 

entity is defined as: (1) A small business 

that is a small industrial entity as 

defined by the Small Business 

Administration’s (SBA) regulations at 13 

CFR 121.201; (2) a small governmental 

jurisdiction that is a government of a 

city, county, town, school district or 

special district with a population of less 

than 50,000; and (3) a small 

organization that is any not-for-profit 

enterprise which is independently 

owned and operated and is not 

dominant in its field. 

After considering the economic 

impacts of today’s proposed rule on 

small entities, I certify that this action 

will not have a significant economic 

impact on a substantial number of small 

entities. This proposed rule will not 

impose any requirements on small 

entities. Rather, this rule establishes 

national standards for allowable 

concentrations of oxides of nitrogen and 

sulfur in ambient air as required by 

section 109 of the CAA. See also 

American Trucking Associations v. 

EPA. 175 F. 3d at 1044–45 (NAAQS do 

not have significant impacts upon small 

entities because NAAQS themselves 

impose no regulations upon small 

entities). We continue to be interested in 

the potential impacts of the proposed 

rule on small entities and welcome 

comments on issues related to such 

impacts. 

D. Unfunded Mandates Reform Act 

Title II of the Unfunded Mandates 

Reform Act of 1995 (UMRA), Public 

Law 104–4, establishes requirements for 

Federal agencies to assess the effects of 

their regulatory actions on State, local, 

and Tribal governments and the private 

sector. Under section 202 of the UMRA, 

EPA generally must prepare a written 

statement, including a cost-benefit 

analysis, for proposed and final rules 

with ‘‘Federal mandates’’ that may 

result in expenditures to state, local, 

and tribal governments, in the aggregate, 

or to the private sector, of $100 million 

or more in any 1 year. Before 

promulgating an EPA rule for which a 

written statement is needed, section 205 

of the UMRA generally requires EPA to 

identify and consider a reasonable 

number of regulatory alternatives and to 

adopt the least costly, most costeffective or least burdensome alternative 

that achieves the objectives of the rule. 

The provisions of section 205 do not 

apply when they are inconsistent with 

applicable law. Moreover, section 205 

allows EPA to adopt an alternative other 

than the least costly, most cost-effective 

or least burdensome alternative if the 

Administrator publishes with the final 

rule an explanation why that alternative 

was not adopted. Before EPA establishes 

any regulatory requirements that may 

significantly or uniquely affect small 

governments, including tribal 

governments, it must have developed 

under section 203 of the UMRA a small 

government agency plan. The plan must 

provide for notifying potentially 

affected small governments, enabling 

officials of affected small governments 

to have meaningful and timely input in 

the development of EPA regulatory 

proposals with significant Federal 

intergovernmental mandates, and 

informing, educating, and advising 

small governments on compliance with 

the regulatory requirements. 

This action contains no Federal 

mandates under the provisions of Title 

II of the Unfunded Mandates Reform 

Act of 1995 (UMRA), 2 U.S.C. 1531– 

1538 for state, local, or tribal 

governments or the private sector. 

Therefore, this action is not subject to 

the requirements of sections 202 or 205. 

Furthermore, as indicated previously, in 

setting a NAAQS EPA cannot consider 

the economic or technological feasibility 

of attaining ambient air quality 

standards; although such factors may be 

considered to a degree in the 

development of state plans to 

implement the standards. See also 

American Trucking Associations v. 

EPA, 175 F. 3d at 1043 (noting that 

because EPA is precluded from 

considering costs of implementation in 

establishing NAAQS, preparation of a 

Regulatory Impact Analysis pursuant to 

the Unfunded Mandates Reform Act 

would not furnish any information 

which the court could consider in 

reviewing the NAAQS). Accordingly, 

EPA has determined that the provisions 

of sections 202, 203, and 205 of the 

UMRA do not apply to this proposed 

decision. The EPA acknowledges, 

however, that any corresponding 

revisions to associated state 

implementation plan (SIP) requirements 

and air quality surveillance 

requirements, 40 CFR part 51 and 40 

CFR part 58, respectively, might result 

in such effects. Accordingly, EPA will 

address, as appropriate, unfunded 

mandates if and when it proposes any 

revisions to 40 CFR parts 51 or 58. 

E. Executive Order 13132: Federalism 

This proposed rule does not have 

federalism implications. It will not have 

substantial direct effects on the states, 

on the relationship between the national 

government and the states, or on the 

distribution of power and 

responsibilities among the various 

levels of government, as specified in 

Executive Order 13132 because it does 

not contain legally binding 

requirements. Thus, the requirements of 

Executive Order 13132 do not apply to 

this rule. 

EPA believes, however, that this 

proposed rule may be of significant 

interest to state governments. As also 

noted in section E (above) on UMRA, 

EPA recognizes that states will have a 

substantial interest in this rule and any 

corresponding revisions to associated 

SIP requirements and air quality 

surveillance requirements, 40 CFR part 

51 and 40 CFR part 58, respectively. 

Therefore, in the spirit of Executive 

Order 13132 and consistent with EPA 

policy to promote communications 

between EPA and state and local 

governments, EPA specifically solicits 

comment on this proposed rule from 

state and local officials. 

F. Executive Order 13175: Consultation 

and Coordination With Indian Tribal 

Governments 

Executive Order 13175, entitled 

‘‘Consultation and Coordination with 

Indian Tribal Governments’’ (65 FR 

67249, November 9, 2000), requires EPA 

to develop an accountable process to 

ensure ‘‘meaningful and timely input by 

tribal officials in the development of 

regulatory policies that have tribal 

implications.’’ This rule concerns the 

establishment of national standards to 

address the public welfare effects of 

oxides of nitrogen and sulfur. 


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This action does not have Tribal 

implications, as specified in Executive 

Order 13175 (65 FR 67249, November 9, 

2000). It does not have a substantial 

direct effect on one or more Indian 

tribes, since tribes are not obligated to 

adopt or implement any NAAQS. Thus, 

Executive Order 13175 does not apply 

to this rule. 

G. Executive Order 13045: Protection of 

Children from Environmental Health & 

Safety Risks 

This action is not subject to EO 13045 

because it is not an economically 

significant rule as defined in EO 12866. 

H. Executive Order 13211: Actions that 

Significantly Affect Energy Supply, 

Distribution or Use 

This action is not a ‘‘significant 

energy action’’ as defined in Executive 

Order 13211 (66 FR 28355, May 22, 

2001), because it is not likely to have a 

significant adverse effect on the supply, 

distribution, or use of energy. This 

action concerns the establishment of 

national standards to address the public 

welfare effects of oxides of nitrogen and 

sulfur. This action does not prescribe 

specific pollution control strategies by 

which these ambient standards will be 

met. Such strategies will be developed 

by states on a case-by-case basis, and 

EPA cannot predict whether the control 

options selected by states will include 

regulations on energy suppliers, 

distributors, or users. 

I. National Technology Transfer and 

Advancement Act 

Section 12(d) of the National 

Technology Transfer and Advancement 

Act of 1995 (NTTAA), Public Law 104– 

113, 12(d) (15 U.S.C. 272 note) directs 

EPA to use voluntary consensus 

standards in its regulatory activities 

unless to do so would be inconsistent 

with applicable law or otherwise 

impractical. Voluntary consensus 

standards are technical standards (e.g., 

materials specifications, test methods, 

sampling procedures, and business 

practices) that are developed or adopted 

by voluntary consensus standards 

bodies. The NTTAA directs EPA to 

provide Congress, through OMB, 

explanations when the Agency decides 

not to use available and applicable 

voluntary consensus standards. 

The EPA is not aware of any 

voluntary consensus standards that are 

relevant to the provisions of this 

proposed rule. The EPA welcomes any 

feedback on such standards that may be 

applicable. 

J. Executive Order 12898: Federal 

Actions To Address Environmental 

Justice in Minority Populations and 

Low-Income Populations 

Executive Order 12898 (59 FR 7629 

(Feb. 16, 1994)) establishes federal 

executive policy on environmental 

justice. Its main provision directs 

federal agencies, to the greatest extent 

practicable and permitted by law, to 

make environmental justice part of their 

mission by identifying and addressing, 

as appropriate, disproportionately high 

and adverse human health or 

environmental effects of their programs, 

policies, and activities on minority 

populations and low-income 

populations in the United States. 

EPA has determined that this 

proposed rule will not have 

disproportionately high and adverse 

human health or environmental effects 

on minority or low-income populations 

because it retains the level of 

environmental protection for all affected 

populations without having any 

disproportionately high and adverse 

human health or environmental effects 

on any population, including any 

minority or low-income population. 

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US EPA, 2011b. Federal Reference 

Methods for NOy and p-SO4 for the New 

Combined NOX and SOx Secondary NAAQS 

Research Plan, EPA/600/1–11/002 January 

20, 2011. 

Wolff, G. T. 1993. CASAC closure letter for 

the 1993 Criteria Document for Oxides of 

Nitrogen addressed to U.S. EPA 

Administrator Carol M. Browner dated 

September 30, 1993. 

Wolff, G. T. 1995. CASAC closure letter for 

the 1995 OAQPS Staff Paper addressed to 

U.S. EPA Administrator Carol M. Browner 

dated August 22, 1995. 

List of Subjects in 40 CFR Part 50 

Environmental protection, Air 

pollution control, Carbon monoxide, 

Lead, Nitrogen dioxide, Ozone, 

Particulate matter, Sulfur oxides. 

Dated: July 12, 2011. 

Lisa P. Jackson, 

Administrator. 

For the reasons set forth in the 

preamble, part 50 of chapter 1 of title 40 

of the code of Federal regulations is 

proposed to be amended as follows: 

PART 50—NATIONAL PRIMARY AND 

SECONDARY AMBIENT AIR QUALITY 

STANDARDS 

1. The authority citation for part 50 

continues to read as follows: 

Authority: 42 U.S.C. 7401, et seq. 

2. Section 50.5 is amended by revising 

paragraphs (b) and (c) and by adding 

paragraphs (d) and (e) to read as follows: 

§ 50.5 National secondary ambient air 

quality standards for sulfur oxides (sulfur 

dioxide). 

* * * * * 

(b) The level of the national secondary 

1-hour ambient air quality standard for 

oxides of sulfur is 75 parts per billion 

(ppb, which is 1 part in 1,000,000,000), 

measured in the ambient air as sulfur 

dioxide (SO2). 

(c) The levels of the standards shall be 

measured by a reference method based 

on Appendix A–1 or A–2 of this part, 

or by a Federal Equivalent Method 

(FEM) designated in accordance with 

part 53 of this chapter. 

(d) To demonstrate attainment with 

the 3-hour secondary standard, the 

second-highest 3-hour average must be 

based upon hourly data that are at least 

75 percent complete in each calendar 

quarter. A 3-hour block average shall be 

considered valid only if all three hourly 

averages for the 3-hour period are 

available. If only one or two hourly 

averages are available, but the 3-hour 

average would exceed the level of the 

standard when zeros are substituted for 

the missing values, subject to the 

rounding rule of paragraph (a) of this 

section, then this shall be considered a 

valid 3-hour average. In all cases, the 3- 

hour block average shall be computed as 

the sum of the hourly averages divided 

by 3. 

(e) The 1-hour secondary standard is 

met at an ambient air quality monitoring 

site when the three-year average of the 

annual 99th percentile of the daily 

maximum 1-hour average 

concentrations is less than or equal to 

75 ppb, as determined in accordance 

with Appendix T of this part. 

3. Section 50.11 is revised to read as 

follows: 

§ 50.11 National primary and secondary 

ambient air quality standards for oxides of 

nitrogen (with nitrogen dioxide as the 

indicator). 

(a) The level of the national primary 

and secondary annual ambient air 

quality standards for oxides of nitrogen 

is 53 parts per billion (ppb, which is 1 

part in 1,000,000,000), annual average 

concentration, measured in the ambient 

air as nitrogen dioxide. 

(b) The level of the national primary 

and secondary 1-hour ambient air 

quality standards for oxides of nitrogen 

is 100 ppb, 1-hour average 

concentration, measured in the ambient 

air as nitrogen dioxide. 

(c) The levels of the standards shall be 

measured by: 

(1) A reference method based on 

appendix F to this part; or 

(2) A Federal equivalent method 

(FEM) designated in accordance with 

part 53 of this chapter. 

(d) The annual primary and secondary 

standards are met when the annual 

average concentration in a calendar year 

is less than or equal to 53 ppb, as 

determined in accordance with 

Appendix S of this part for the annual 

standard. 

(e) The 1-hour primary and secondary 

standards are met when the three-year 

average of the annual 98th percentile of 

the daily maximum 1-hour average 

concentration is less than or equal to 

100 ppb, as determined in accordance 

with Appendix S of this part for the 1- 

hour standard. 

4. Appendix S is amended as follows: 

a. by revising paragraph 1.(a), 

b. by revising the definition of 

‘‘Design values’’ under paragraph 1.(c), 

c. by revising paragraph 2.(b), 

d. by revising paragraphs 3.1(a) 

through (d), 

e. by revising paragraphs 3.2(a) 

through (e), 

f. by revising paragraph 4.1(b), 

g. by revising paragraph 4.2(c), 

h. by revising paragraph 5.1(b), and 

i. by revising paragraph 5.2(b) to read 

as follows: 


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Federal Register / Vol. 76, No. 147 / Monday, August 1, 2011 / Proposed Rules 46145 

Appendix S to Part 50—Interpretation 

of the Primary and Secondary National 

Ambient Air Quality Standards for 

Oxides of Nitrogen (Nitrogen Dioxide) 

1. General. 

(a) This appendix explains the data 

handling conventions and computations 

necessary for determining when the primary 

and secondary national ambient air quality 

standards for oxides of nitrogen as measured 

by nitrogen dioxide (‘‘NO2 NAAQS’’) 

specified in § 50.11 are met. Nitrogen dioxide 

(NO2) is measured in the ambient air by a 

Federal reference method (FRM) based on 

appendix F to this part or by a Federal 

equivalent method (FEM) designated in 

accordance with part 53 of this chapter. Data 

handling and computation procedures to be 

used in making comparisons between 

reported NO2 concentrations and the levels of 

the NO2 NAAQS are specified in the 

following sections. 

* * * * * 

(c) * * * 

Design values are the metrics (i.e., 

statistics) that are compared to the NAAQS 

levels to determine compliance, calculated as 

specified in section 5 of this appendix. The 

design values for the primary and secondary 

NAAQS are: 

(1) The annual mean value for a monitoring 

site for one year (referred to as the ‘‘annual 

primary or secondary standard design 

value’’). 

(2) The 3-year average of annual 98th 

percentile daily maximum 1-hour values for 

a monitoring site (referred to as the ‘‘1-hour 

primary or secondary standard design 

value’’). 

* * * * * 

2. Requirements for Data Used for 

Comparisons With the NO2 NAAQS and Data 

Reporting Considerations. 

* * * * * 

(b) When two or more NO2 monitors are 

operated at a site, the state may in advance 

designate one of them as the primary 

monitor. If the state has not made this 

designation, the Administrator will make the 

designation, either in advance or 

retrospectively. Design values will be 

developed using only the data from the 

primary monitor, if this results in a valid 

design value. If data from the primary 

monitor do not allow the development of a 

valid design value, data solely from the other 

monitor(s) will be used in turn to develop a 

valid design value, if this results in a valid 

design value. If there are three or more 

monitors, the order for such comparison of 

the other monitors will be determined by the 

Administrator. The Administrator may 

combine data from different monitors in 

different years for the purpose of developing 

a valid 1-hour primary or secondary standard 

design value, if a valid design value cannot 

be developed solely with the data from a 

single monitor. However, data from two or 

more monitors in the same year at the same 

site will not be combined in an attempt to 

meet data completeness requirements, except 

if one monitor has physically replaced 

another instrument permanently, in which 

case the two instruments will be considered 

to be the same monitor, or if the state has 

switched the designation of the primary 

monitor from one instrument to another 

during the year. 

* * * * * 

3. Comparisons with the NO2 NAAQS. 

3.1 The Annual Primary and Secondary 

NO2 NAAQS. 

(a) The annual primary and secondary NO2

NAAQS are met at a site when the valid 

annual primary standard design value is less 

than or equal to 53 parts per billion (ppb). 

(b) An annual primary or secondary 

standard design value is valid when at least 

75 percent of the hours in the year are 

reported. 

(c) An annual primary or secondary 

standard design value based on data that do 

not meet the completeness criteria stated in 

section 3.1(b) may also be considered valid 

with the approval of, or at the initiative of, 

the Administrator, who may consider factors 

such as monitoring site closures/moves, 

monitoring diligence, the consistency and 

levels of the valid concentration 

measurements that are available, and nearby 

concentrations in determining whether to use 

such data. 

(d) The procedures for calculating the 

annual primary and secondary standard 

design values are given in section 5.1 of this 

appendix. 

3.2 The 1-Hour Primary and Secondary 

NO2 NAAQS. 

(a) The 1-hour primary or secondary NO2

NAAQS is met at a site when the valid 1- 

hour primary or secondary standard design 

value is less than or equal to 100 parts per 

billion (ppb). 

(b) An NO2 1-hour primary or secondary 

standard design value is valid if it 

encompasses three consecutive calendar 

years of complete data. A year meets data 

completeness requirements when all 4 

quarters are complete. A quarter is complete 

when at least 75 percent of the sampling days 

for each quarter have complete data. A 

sampling day has complete data if 75 percent 

of the hourly concentration values, including 

state-flagged data affected by exceptional 

events which have been approved for 

exclusion by the Administrator, are reported. 

(c) In the case of one, two, or three years 

that do not meet the completeness 

requirements of section 3.2(b) of this 

appendix and thus would normally not be 

useable for the calculation of a valid 3-year 

1-hour primary or secondary standard design 

value, the 3-year 1-hour primary or 

secondary standard design value shall 

nevertheless be considered valid if one of the 

following conditions is true. 

(i) At least 75 percent of the days in each 

quarter of each of three consecutive years 

have at least one reported hourly value, and 

the design value calculated according to the 

procedures specified in section 5.2 is above 

the level of the primary or secondary 1-hour 

standard. 

(ii) (A) A 1-hour primary or secondary 

standard design value that is below the level 

of the NAAQS can be validated if the 

substitution test in section 3.2(c)(ii)(B) results 

in a ‘‘test design value’’ that is below the 

level of the NAAQS. The test substitutes 

actual ‘‘high’’ reported daily maximum 1- 

hour values from the same site at about the 

same time of the year (specifically, in the 

same calendar quarter) for unknown values 

that were not successfully measured. Note 

that the test is merely diagnostic in nature, 

intended to confirm that there is a very high 

likelihood that the original design value (the 

one with less than 75 percent data capture of 

hours by day and of days by quarter) reflects 

the true under-NAAQS-level status for that 3- 

year period; the result of this data 

substitution test (the ‘‘test design value’’, as 

defined in section 3.2(c)(ii)(B)) is not 

considered the actual design value. For this 

test, substitution is permitted only if there 

are at least 200 days across the three 

matching quarters of the three years under 

consideration (which is about 75 percent of 

all possible daily values in those three 

quarters) for which 75 percent of the hours 

in the day, including state-flagged data 

affected by exceptional events which have 

been approved for exclusion by the 

Administrator, have reported concentrations. 

However, maximum 1-hour values from days 

with less than 75 percent of the hours 

reported shall also be considered in 

identifying the high value to be used for 

substitution. 

(B) The substitution test is as follows: Data 

substitution will be performed in all quarter 

periods that have less than 75 percent data 

capture but at least 50 percent data capture, 

including state-flagged data affected by 

exceptional events which have been 

approved for exclusion by the Administrator; 

if any quarter has less than 50 percent data 

capture then this substitution test cannot be 

used. Identify for each quarter (e.g., January– 

March) the highest reported daily maximum 

1-hour value for that quarter, excluding stateflagged data affected by exceptional events 

which have been approved for exclusion by 

the Administrator, looking across those three 

months of all three years under 

consideration. All daily maximum 1-hour 

values from all days in the quarter period 

shall be considered when identifying this 

highest value, including days with less than 

75 percent data capture. If after substituting 

the highest non-excluded reported daily 

maximum 1-hour value for a quarter for as 

much of the missing daily data in the 

matching deficient quarter(s) as is needed to 

make them 100 percent complete, the 

procedure in section 5.2 yields a recalculated 

3-year 1-hour standard ‘‘test design value’’ 

below the level of the standard, then the 1- 

hour primary or secondary standard design 

value is deemed to have passed the 

diagnostic test and is valid, and the level of 

the standard is deemed to have been met in 

that 3-year period. As noted in section 

3.2(c)(i), in such a case, the 3-year design 

value based on the data actually reported, not 

the ‘‘test design value’’, shall be used as the 

valid design value. (iii) (A) A 1-hour primary 

or secondary standard design value that is 

above the level of the NAAQS can be 

validated if the substitution test in section 

3.2(c)(iii)(B) results in a ‘‘test design value’’ 

that is above the level of the NAAQS. The 

test substitutes actual ‘‘low’’ reported daily 

maximum 1-hour values from the same site 

at about the same time of the year 

(specifically, in the same three months of the 


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46146 Federal Register / Vol. 76, No. 147 / Monday, August 1, 2011 / Proposed Rules 

calendar) for unknown values that were not 

successfully measured. Note that the test is 

merely diagnostic in nature, intended to 

confirm that there is a very high likelihood 

that the original design value (the one with 

less than 75 percent data capture of hours by 

day and of days by quarter) reflects the true 

above-NAAQS-level status for that 3-year 

period; the result of this data substitution test 

(the ‘‘test design value’’, as defined in section 

3.2(c)(iii)(B)) is not considered the actual 

design value. For this test, substitution is 

permitted only if there are a minimum 

number of available daily data points from 

which to identify the low quarter-specific 

daily maximum 1-hour values, specifically if 

there are at least 200 days across the three 

matching quarters of the three years under 

consideration (which is about 75 percent of 

all possible daily values in those three 

quarters) for which 75 percent of the hours 

in the day have reported concentrations. 

Only days with at least 75 percent of the 

hours reported shall be considered in 

identifying the low value to be used for 

substitution. 

(B) The substitution test is as follows: Data 

substitution will be performed in all quarter 

periods that have less than 75 percent data 

capture. Identify for each quarter (e.g., 

January–March) the lowest reported daily 

maximum 1-hour value for that quarter, 

looking across those three months of all three 

years under consideration. All daily 

maximum 1-hour values from all days with 

at least 75 percent capture in the quarter 

period shall be considered when identifying 

this lowest value. If after substituting the 

lowest reported daily maximum 1-hour value 

for a quarter for as much of the missing daily 

data in the matching deficient quarter(s) as is 

needed to make them 75 percent complete, 

the procedure in section 5.2 yields a 

recalculated 3-year 1-hour standard ‘‘test 

design value’’ above the level of the standard, 

then the 1-hour primary or secondary 

standard design value is deemed to have 

passed the diagnostic test and is valid, and 

the level of the standard is deemed to have 

been exceeded in that 3-year period. As 

noted in section 3.2(c)(i), in such a case, the 

3-year design value based on the data 

actually reported, not the ‘‘test design value’’, 

shall be used as the valid design value. 

(d) A 1-hour primary or secondary 

standard design value based on data that do 

not meet the completeness criteria stated in 

3.2(b) and also do not satisfy section 3.2(c), 

may also be considered valid with the 

approval of, or at the initiative of, the 

Administrator, who may consider factors 

such as monitoring site closures/moves, 

monitoring diligence, the consistency and 

levels of the valid concentration 

measurements that are available, and nearby 

concentrations in determining whether to use 

such data. 

(e) The procedures for calculating the 1- 

hour primary and secondary standard design 

values are given in section 5.2 of this 

appendix. 

4. Rounding Conventions. 

4.1 Rounding Conventions for the Annual 

Primary and Secondary NO2 NAAQS. 

* * * * * 

(b) The annual primary or secondary 

standard design value is calculated pursuant 

to section 5.1 and then rounded to the 

nearest whole number or 1 ppb (decimals 0.5 

and greater are rounded up to the nearest 

whole number, and any decimal lower than 

0.5 is rounded down to the nearest whole 

number). 

4.2 Rounding Conventions for the 1-hour 

Primary and Secondary NO2 NAAQS. 

* * * * * 

(c) The 1-hour primary or secondary 

standard design value is calculated pursuant 

to section 5.2 and then rounded to the 

nearest whole number or 1 ppb (decimals 0.5 

and greater are rounded up to the nearest 

whole number, and any decimal lower than 

0.5 is rounded down to the nearest whole 

number). 

5. Calculation Procedures for the Primary 

and Secondary NO2 NAAQS. 

5.1 Procedures for the Annual Primary 

and Secondary NO2 NAAQS. 

* * * * * 

(b) The annual primary or secondary 

standard design value for a site is the valid 

annual mean rounded according to the 

conventions in section 4.1. 

5.2 Calculation Procedures for the 1-hour 

Primary and Secondary NO2 NAAQS. 

* * * * * 

(b) The 1-hour primary or secondary 

standard design value for a site is the mean 

of the three annual 98th percentile values, 

rounded according to the conventions in 

section 4. 

* * * * * 

5. Appendix T is amended as follows: 

a. by revising paragraph 1.(a), 

b. by revising the definition of 

‘‘Design values’’ under paragraph 1.(c), 

c. by revising paragraph 2.(b), 

d. by revising paragraphs 3.(a) 

through (e), 

e. by revising paragraph 4.(c), and 

f. by revising paragraph 5.(b) to read 

as follows: 

Appendix T to Part 50—Interpretation 

of the Primary and Secondary National 

Ambient Air Quality Standards for 

Oxides of Sulfur (Sulfur Dioxide) 

1. General. 

(a) This appendix explains the data 

handling conventions and computations 

necessary for determining when the primary 

and secondary national ambient air quality 

standards for Oxides of Sulfur as measured 

by Sulfur Dioxide (‘‘SO2 NAAQS’’) specified 

in § 50.17 and § 50.5 (b), respectively, are met 

at an ambient air quality monitoring site. 

Sulfur dioxide (SO2) is measured in the 

ambient air by a Federal reference method 

(FRM) based on appendix A–1 or A–2 to this 

part or by a Federal equivalent method (FEM) 

designated in accordance with part 53 of this 

chapter. Data handling and computation 

procedures to be used in making 

comparisons between reported SO2

concentrations and the levels of the SO2

NAAQS are specified in the following 

sections. 

* * * * * 

(c) * * * 

Design values are the metrics (i.e., 

statistics) that are compared to the NAAQS 

levels to determine compliance, calculated as 

specified in section 5 of this appendix. The 

design value for the primary and secondary 

1-hour NAAQS is the 3-year average of 

annual 99th percentile daily maximum 1- 

hour values for a monitoring site (referred to 

as the ‘‘1-hour primary standard design 

value’’). 

* * * * * 

2. Requirements for Data Used for 

Comparisons With the SO2 NAAQS and Data 

Reporting Considerations. 

* * * * * 

(b) Data from two or more monitors from 

the same year at the same site reported to 

EPA under distinct Pollutant Occurrence 

Codes shall not be combined in an attempt 

to meet data completeness requirements. The 

Administrator will combine annual 99th 

percentile daily maximum concentration 

values from different monitors in different 

years, selected as described here, for the 

purpose of developing a valid 1-hour primary 

or secondary standard design value. If more 

than one of the monitors meets the 

completeness requirement for all four 

quarters of a year, the steps specified in 

section 5(a) of this appendix shall be applied 

to the data from the monitor with the highest 

average of the four quarterly completeness 

values to derive a valid annual 99th 

percentile daily maximum concentration. If 

no monitor is complete for all four quarters 

in a year, the steps specified in section 3(c) 

and 5(a) of this appendix shall be applied to 

the data from the monitor with the highest 

average of the four quarterly completeness 

values in an attempt to derive a valid annual 

99th percentile daily maximum 

concentration. This paragraph does not 

prohibit a monitoring agency from making a 

local designation of one physical monitor as 

the primary monitor for a Pollutant 

Occurrence Code and substituting the 1-hour 

data from a second physical monitor 

whenever a valid concentration value is not 

obtained from the primary monitor; if a 

monitoring agency substitutes data in this 

manner, each substituted value must be 

accompanied by an AQS qualifier code 

indicating that substitution with a value from 

a second physical monitor has taken place. 

* * * * * 

3. Comparisons with the 1-hour Primary 

and Secondary SO2 NAAQS. 

(a) The 1-hour primary or secondary SO2

NAAQS is met at an ambient air quality 

monitoring site when the valid 1-hour 

primary or secondary standard design value 

is less than or equal to 75 parts per billion 

(ppb). 

(b) An SO2 1-hour primary or secondary 

standard design value is valid if it 

encompasses three consecutive calendar 

years of complete data. A year meets data 

completeness requirements when all 4 

quarters are complete. A quarter is complete 

when at least 75 percent of the sampling days 

for each quarter have complete data. A 

sampling day has complete data if 75 percent 

of the hourly concentration values, including 

State-flagged data affected by exceptional 

events which have been approved for 

exclusion by the Administrator, are reported. 


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(c) In the case of one, two, or three years 

that do not meet the completeness 

requirements of section 3(b) of this appendix 

and thus would normally not be useable for 

the calculation of a valid 3-year 1-hour 

primary or secondary standard design value, 

the 3-year 1-hour primary or secondary 

standard design value shall nevertheless be 

considered valid if one of the following 

conditions is true. 

(i) At least 75 percent of the days in each 

quarter of each of three consecutive years 

have at least one reported hourly value, and 

the design value calculated according to the 

procedures specified in section 5 is above the 

level of the primary or secondary 1-hour 

standard. 

(ii) (A) A 1-hour primary or secondary 

standard design value that is equal to or 

below the level of the NAAQS can be 

validated if the substitution test in section 

3(c)(ii)(B) results in a ‘‘test design value’’ that 

is below the level of the NAAQS. The test 

substitutes actual ‘‘high’’ reported daily 

maximum 1-hour values from the same site 

at about the same time of the year 

(specifically, in the same calendar quarter) 

for unknown values that were not 

successfully measured. Note that the test is 

merely diagnostic in nature, intended to 

confirm that there is a very high likelihood 

that the original design value (the one with 

less than 75 percent data capture of hours by 

day and of days by quarter) reflects the true 

under-NAAQS-level status for that 3-year 

period; the result of this data substitution test 

(the ‘‘test design value’’, as defined in section 

3(c)(ii)(B)) is not considered the actual design 

value. For this test, substitution is permitted 

only if there are at least 200 days across the 

three matching quarters of the three years 

under consideration (which is about 75 

percent of all possible daily values in those 

three quarters) for which 75 percent of the 

hours in the day, including State-flagged data 

affected by exceptional events which have 

been approved for exclusion by the 

Administrator, have reported concentrations. 

However, maximum 1-hour values from days 

with less than 75 percent of the hours 

reported shall also be considered in 

identifying the high value to be used for 

substitution. 

(B) The substitution test is as follows: Data 

substitution will be performed in all quarter 

periods that have less than 75 percent data 

capture but at least 50 percent data capture, 

including State-flagged data affected by 

exceptional events which have been 

approved for exclusion by the Administrator; 

if any quarter has less than 50 percent data 

capture then this substitution test cannot be 

used. Identify for each quarter (e.g., January– 

March) the highest reported daily maximum 

1-hour value for that quarter, excluding Stateflagged data affected by exceptional events 

which have been approved for exclusion by 

the Administrator, looking across those three 

months of all three years under 

consideration. All daily maximum 1-hour 

values from all days in the quarter period 

shall be considered when identifying this 

highest value, including days with less than 

75 percent data capture. If after substituting 

the highest reported daily maximum 1-hour 

value for a quarter for as much of the missing 

daily data in the matching deficient 

quarter(s) as is needed to make them 100 

percent complete, the procedure in section 5 

yields a recalculated 3-year 1-hour standard 

‘‘test design value’’ less than or equal to the 

level of the standard, then the 1-hour primary 

or secondary standard design value is 

deemed to have passed the diagnostic test 

and is valid, and the level of the standard is 

deemed to have been met in that 3-year 

period. As noted in section 3(c)(i), in such a 

case, the 3-year design value based on the 

data actually reported, not the ‘‘test design 

value’’, shall be used as the valid design 

value. 

(iii) (A) A 1-hour primary or secondary 

standard design value that is above the level 

of the NAAQS can be validated if the 

substitution test in section 3(c)(iii)(B) results 

in a ‘‘test design value’’ that is above the 

level of the NAAQS. The test substitutes 

actual ‘‘low’’ reported daily maximum 1-hour 

values from the same site at about the same 

time of the year (specifically, in the same 

three months of the calendar) for unknown 

hourly values that were not successfully 

measured. Note that the test is merely 

diagnostic in nature, intended to confirm that 

there is a very high likelihood that the 

original design value (the one with less than 

75 percent data capture of hours by day and 

of days by quarter) reflects the true aboveNAAQS-level status for that 3-year period; 

the result of this data substitution test (the 

‘‘test design value’’, as defined in section 

3(c)(iii)(B)) is not considered the actual 

design value. For this test, substitution is 

permitted only if there are a minimum 

number of available daily data points from 

which to identify the low quarter-specific 

daily maximum 1-hour values, specifically if 

there are at least 200 days across the three 

matching quarters of the three years under 

consideration (which is about 75 percent of 

all possible daily values in those three 

quarters) for which 75 percent of the hours 

in the day have reported concentrations. 

Only days with at least 75 percent of the 

hours reported shall be considered in 

identifying the low value to be used for 

substitution. 

(B) The substitution test is as follows: Data 

substitution will be performed in all quarter 

periods that have less than 75 percent data 

capture. Identify for each quarter (e.g., 

January–March) the lowest reported daily 

maximum 1-hour value for that quarter, 

looking across those three months of all three 

years under consideration. All daily 

maximum 1-hour values from all days with 

at least 75 percent capture in the quarter 

period shall be considered when identifying 

this lowest value. If after substituting the 

lowest reported daily maximum 1-hour value 

for a quarter for as much of the missing daily 

data in the matching deficient quarter(s) as is 

needed to make them 75 percent complete, 

the procedure in section 5 yields a 

recalculated 3-year 1-hour standard ‘‘test 

design value’’ above the level of the standard, 

then the 1-hour primary or secondary 

standard design value is deemed to have 

passed the diagnostic test and is valid, and 

the level of the standard is deemed to have 

been exceeded in that 3-year period. As 

noted in section 3(c)(i), in such a case, the 

3-year design value based on the data 

actually reported, not the ‘‘test design value’’, 

shall be used as the valid design value. 

(d) A 1-hour primary or secondary 

standard design value based on data that do 

not meet the completeness criteria stated in 

3(b) and also do not satisfy section 3(c), may 

also be considered valid with the approval of, 

or at the initiative of, the Administrator, who 

may consider factors such as monitoring site 

closures/moves, monitoring diligence, the 

consistency and levels of the valid 

concentration measurements that are 

available, and nearby concentrations in 

determining whether to use such data. 

(e) The procedures for calculating the 1- 

hour primary or secondary standard design 

values are given in section 5 of this 

appendix. 

4. Rounding Conventions for the 1-hour 

Primary and Secondary SO2 NAAQS. 

* * * * * 

(c) The 1-hour primary or secondary 

standard design value is calculated pursuant 

to section 5 and then rounded to the nearest 

whole number or 1 ppb (decimals 0.5 and 

greater are rounded up to the nearest whole 

number, and any decimal lower than 0.5 is 

rounded down to the nearest whole number). 

5. Calculation Procedures for the 1-hour 

Primary and Secondary SO2 NAAQS. 

* * * * * 

(b) The 1-hour primary or secondary 

standard design value for an ambient air 

quality monitoring site is the mean of the 

three annual 99th percentile values, rounded 

according to the conventions in section 4. 

[FR Doc. 2011–18582 Filed 7–29–11; 8:45 am] 

BILLING CODE 6560–50–P 


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