Policy Assessment for the Review of the 

Secondary National Ambient Air Quality 

Standards for NOx and SOx 

Second External Review Draft 
 EPA 452/P-10-008 

September, 2010 

Policy Assessment for the Review of the 

Secondary National Ambient Air Quality Standards 

for NOx and SOx: 

Second External Review Draft

U.S. Environmental Protection Agency 

Office of Air Quality Planning and Standards 

Health and Environmental Impacts Division 

Research Triangle Park, North Carolina 
DISCLAIMER

This document has been reviewed by the Office of Air Quality Planning and Standards, 

U.S. Environmental Protection Agency (EPA), and approved for publication. This draft 

document has been prepared by staff from the Office of Air Quality Planning and Standards, 

U.S. Environmental Protection Agency. Any opinions, findings, conclusions, or 

recommendations are those of the authors and do not necessarily reflect the views of the EPA 

Mention of trade names or commercial products is not intended to constitute endorsement or 

recommendation for use. This document is being provided to the Clean Air Scientific Advisory 

Committee for their review, and made available to the public for comment. Any questions or 

comments concerning this document should be addressed to Bryan Hubbell, U.S. Environmental 

Protection Agency, Office of Air Quality Planning and Standards, C504-06, Research Triangle 

Park, North Carolina 27711 (email: hubbell.bryan@epa.gov ). 

Acknowledgements (if any) 
Policy Assessment for the Review of the Secondary National Ambient Air 

Quality Standards for NOx and SOx: Second External Review Draft 

EXECUTIVE SUMMARY 

Introduction 

This second draft Policy Assessment is an evaluation of the policy implications of the key 

scientific information contained in the Integrated Science Assessment (ISA) for Oxides of Nitrogen and 

Sulfur-Ecological Criteria, prepared by EPA’s National Center for Environmental Assessment (NCEA), 

and the results from the analyses contained in the Risk and Exposure Assessment (REA) for Review of 

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

This second draft also presents preliminary EPA staff conclusions regarding the adequacy of the current 

standards and various policy options that we believe are appropriate to consider as part of the current 

review of the secondary (welfare-based, e.g. focused on non-health effects including impacts on soils, 

water, crops, vegetation, man-made materials, animals, wildlife, weather, visibility, and climate) NOx and 

SOx NAAQS. 

 

This policy assessment is intended to help “bridge the gap” between the scientific assessment 

contained in the ISA and the judgments required of the EPA Administrator in determining whether, and if 

so, how, it is appropriate to revise the secondary NAAQS for NOx and SOx. This policy assessment 

considers the available scientific evidence and quantitative risk-based analyses, together with related 

limitations and uncertainties, and focuses on the basic elements of air quality standards: indicators, 

averaging times, forms, and levels. These elements, which serve to define each standard, must be 

considered collectively in evaluating the welfare protection afforded by the secondary NOx and SOx

NAAQS. 

In conducting this periodic review of the NOx and SOx secondary NAAQS, EPA has decided to 

jointly assess the scientific information, associated risks, and standards because ambient NOx and SOx, 

and their associated transformation products, such as deposited N and S, are linked from an atmospheric 

chemistry perspective, as well as jointly contributing to environmental effects. 

Scope 

This assessment primarily focuses on the effects of the deposition of ambient NOx and SOx on 

multiple ecological receptors. Highlighted effects include those associated with acidification and nitrogen 

nutrient enrichment. Based on these highlighted effects, EPA’s objective is to develop a framework for 

setting standards that are ecologically relevant and that reflect the common impacts of these two 

pollutants as they deposit to sensitive ecosystems. 

For this second draft policy assessment, we have chosen to focus much of our discussion on 

effects in sensitive aquatic ecosystems caused by acidifying deposition of nitrogen and sulfur, which is a 

transformation product of ambient NOx and SOx. We have a high degree of confidence in the link to 

aquatic acidification effects as well as more information available with which to develop an ecologically 

meaningful structure for the standards. We recognize in doing so that the resulting standards will not 

likely provide full protection against terrestrial acidification effects or against adverse nutrient enrichment 

effects in sensitive terrestrial and aquatic ecosystems. It is however likely that some additional protection 

for those ecosystems will be provided as overall NOx and SOx levels in the environment are decreased in 

response to the aquatic acidification based standards. 
Draft Policy Assessment for the Review of the Secondary NAAQS for NOx and SOx: 

Executive Summary 

 

In the atmospheric science community, NOx typically refers to the sum of nitrogen dioxide (NO2), 

and nitric oxide (NO). In contrast, the Clean Air Act uses “NOx” to refer to any gaseous mixture of 

species composed solely of nitrogen and oxygen (e.g., NO2, NO, nitrous oxide [N2O], nitrogen trioxide 

[N2O3], nitrogen tetroxide [N2O4], and dinitrogen pentoxide [N2O5]). The term used by the scientific 

community to represent the complete set of oxidized nitrogen compounds, including those listed in CAA 

Section 108(c), is total oxidized nitrogen (NOy). NOy includes all nitrogen oxides, including NO, NO2, 

HNO3, peroxyacetyl nitrate (PAN), 2N2O5, HONO, NO3, organic nitrates, and particulate NO3. In the 

policy assessment, unless otherwise indicated, we use the term “NOy” to refer to the complete set of 

oxidized nitrogen compounds. 

For this assessment, the full definition of SOx includes all oxides of sulfur, including both 

gaseous substances (e.g., SO2, sulfur monoxide [SO], sulfur trioxide [SO3], thiosulfate [S2O3], and 

heptoxide [S2O7], as well as particulate species, such as ammonium sulfate [(NH4)2SO4]). However, 

throughout this document we refer to SOx as the sum of SO2 and sulfate to be consistent with standard 

monitoring instrumentation. Sulfate is referred to as SO4 and nitrate as NO3, recognizing that they refer to 

the ions that have charges of -2 for sulfate and -1 for nitrate. 

Conceptual Framework 

The figure below depicts the framework by which we are considering the structure of an 

ecologically relevant secondary standard. It is a conceptual diagram that illustrates how a level of 

protection related to an indicator of ecological effect(s) equates to atmospheric concentrations of NOx and 

SOx indicators. This conceptual diagram illustrates the linkages between ambient air concentrations and 

resulting deposition metrics, and between the deposition metric and the ecological indicator of concern. 

The Deposition Transference Ratios translate between NOx and SOx deposition metrics and ambient 

atmospheric concentrations of NOx and SOx, while the Ecological Response to Deposition Function 

relates the deposition metric into the ecological indicator. 

Air Quality 

Indicator(s)

Ecological 

Indicator

Ecological 

Response 

to 

Deposition

Function

Atmospheric 

Deposition

Transference

Ratio

Deposition

Location 

specific 

Modifying 

Factors

Location 

specific 

Modifying 

Factors

Air Quality 

Indicator(s)

Ecological 

Indicator

Ecological 

Response 

to 

Deposition

Function

Atmospheric 

Deposition

Transference

Ratio

Deposition

Location 

specific 

Modifying 

Factors

Location 

specific 

Modifying 

Factors

Our policy assessment is structured around this conceptual model, and includes an evaluation of the 

effects associated with deposition of NOx and SOx to ecosystems, as well as an assessment of the 

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Draft Policy Assessment for the Review of the Secondary NAAQS for NOx and SOx: 

Executive Summary 

adequacy of the existing NOx and SOx standards in protecting against these effects. This policy 

assessment also develops a more complete understanding of the conceptual structure needed to address 

the variable ecosystem and atmospheric factors which modify the impacts of deposited NOx and SOx on 

ecosystems. Development of the form for the standard and options for ambient atmospheric indicators for 

NOx and SOx, averaging times, and levels of the standard are also discussed. 

Ecological Effects from NOx and SOx Deposition 

Effects are broadly categorized into those related to acidification and nutrient-enrichment. Acidification 

occurs in both aquatic and terrestrial ecosystems, with most aquatic effects occurring in freshwater lakes 

and streams. Nutrient enrichment also occurs in both aquatic and terrestrial ecosystems; however, the 

types and prevalence of nutrient enrichment effects vary between freshwater and estuarine aquatic 

ecosystems. 

In the process of acidification, chemical components of terrestrial and freshwater aquatic ecosystems are 

altered in a way that leads to effects on biological organisms. Because NOx and SOx deposited to 

terrestrial ecosystems often move through the soil and eventually leach into adjacent water bodies, 

deposition to terrestrial ecosystems is also a cause of acidification in aquatic ecosystems. 

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

effects on biogeochemical processes and biota in aquatic ecosystems, and between acidifying deposition 

and changes in biogeochemistry in terrestrial ecosystems. Acidic deposition is observed to alter sulfate 

and nitrate concentrations in surface waters, balance of base cations, acid neutralizing capacity (ANC), 

inorganic aluminum, calcium, and surface water pH. These changes can result in the loss of acidsensitive biological species such as salmonid fish species and disrupt food web dynamics causing 

alteration to the diet, breeding distribution and reproduction of certain species of bird, such as goldeneye 

ducks and loons. Acidification in terrestrial ecosystems has been shown to cause decreased growth and 

increased susceptibility to disease and injury in sensitive tree species, including red spruce and sugar 

maple. 

Principal factors governing the sensitivity of terrestrial and aquatic ecosystems to acidification from 

sulfur and nitrogen deposition include geology, biological uptake of nitrogen, soil depth, and elevation. 

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 hydrology. Episodic and chronic 

acidification tends to occur primarily (but not exclusively) at relatively high elevations in areas that have 

base-poor bedrock, high relief, and shallow soils. 

Based on published analyses of surface water data from freshwater ecosystem surveys and monitoring, 

the most sensitive lakes and streams are located in New England, the Adirondack Mountains, the 

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

and high elevation Western ecosystems. 

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. Annual or multiyear average ANC is a good overall indicator of sensitivity, capturing the ability 

of an ecosystem to withstand episodic events such as spring melting that can lower ANC over shorter 

time spans. Biota are generally not harmed when annual average ANC levels are >100 microequivalents 

per liter (μeq/L). At annual average ANC levels between 100 and 50 μeq/L, the fitness of sensitive 

species (e.g., brook trout, zooplankton) begins to decline. When annual average ANC is <50 μeq/L, 

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Draft Policy Assessment for the Review of the Secondary NAAQS for NOx and SOx: 

Executive Summary 

negative effects on aquatic biota are observed, including large reductions in diversity of fish species, and 

declines in health of fish populations, affecting reproductive ability and fitness. Annual average ANC 

levels below 50 μeq/L are generally associated with death or loss of fitness of biota that are sensitive to 

acidification. 

 

Recent studies indicate that acidification of lakes and streams can result in significant loss in economic 

value, which is one indicator of adversity associated with loss of ecosystem services. A 2006 study of 

New York residents found that they are willing to pay between $300 and $800 million annually for the 

equivalent of improving lakes in the Adirondacks region to an ANC of 50. In addition, 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. 

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. A commonly used indicator of terrestrial acidification is the base cation to aluminum ratio, 

Bc/Al. Many locations in sensitive areas of the U.S. have Bc/Al levels below benchmark levels we have 

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 reduced by 20 percent. While not defining whether 

a 20 percent reduction in growth can be considered significant, existing economic studies suggest that 

avoiding significant declines in the health of spruce and sugar maple forests may be worth billions of 

dollars to residents of the Eastern U.S. 

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

deposition. 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 N, may result in imbalance in 

ecosystems, with effects on ecosystem processes, structure and function. In certain N-limited ecosystems, 

including many ecosystems managed for commercial production, N deposition can result in beneficial 

increases in productivity. Nutrient enrichment effects from NOx deposition are difficult to disentangle 

from overall effects of nitrogen enrichment. This is caused by two factors: the inputs of reduced nitrogen 

from deposition and, in estuarine ecosystems, a large fraction of nitrogen inputs from non-atmospheric 

sources. 

Adequacy of the Existing NOx and SOx Standards to Protect Against 

Acidification and Nutrient Enrichment Effects 

Current NOx and SOx secondary standards are designed to protect against direct exposure of vegetation to 

ambient concentrations of NOx and SOx. Almost all areas of the U.S. are in attainment of the current NOx

and SOx secondary standards. The NOx secondary standard is 0.053 parts per million (ppm), annual 

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

secondary standard, which uses SO2 as the atmospheric indicator, is a 3-hour average of 0.5 ppm, not to 

be exceeded more than once per year. 

Recent acidification status of aquatic ecosystems indicate that in the Adirondacks and Shenandoah areas, 

rates of acidifying deposition of NOx and SOx are still well above pre-acidification (1860) conditions. 

Forty-four percent of Adirondack lakes evaluated exceed the critical load for an ANC of 50 μeq/L, and in 

these lakes recreationally important fish species such as trout are missing due to acidification. In the 

Shenandoah area, 85 percent of streams evaluated exceed the critical load for an ANC of 50 μeq/L, 

resulting in losses in fitness in species such as the Blacknose Dace. 

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Draft Policy Assessment for the Review of the Secondary NAAQS for NOx and SOx: 

Executive Summary 

The REA only evaluated a small number of sensitive areas as case studies. However, in the sugar maple 

case study area (Kane Experimental Forest, Pennsylvania), recent (2002) deposition levels are associated 

with a Bc/Al ratio below 1.2, indicating the potential for a greater than 20 percent reduction in growth. In 

the red spruce case study area (Hubbard Brook Experimental Forest, New Hampshire), recent deposition 

levels are associated with a Bc/Al ratio slightly above 1.2, indicating slightly less potential for significant 

reductions in growth. 

Available ecological indicators for estuarine nutrient enrichment are not sufficiently sensitive to changes 

in atmospheric NOx to be of use in assessing the adequacy of existing NOx standards. Atmospheric NOx

can be an important contributor of N to estuarine nutrient enrichment, but additional analysis is required 

to develop an appropriate indicator for assessing levels of protection from nutrient enrichment effects in 

estuaries related to deposition of NOx. 

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 within the range 

associated with ambient NOx levels meeting the current standards, are known to cause changes in species 

composition in diatom communities indicating impaired water quality. From this we initially conclude 

that the existing secondary standard for NOx does not protect such ecosystems and their resulting services 

from impairment. 

Most terrestrial ecosystems in the US are N-limited, and therefore they are sensitive to perturbation 

caused by N additions. Under recent conditions, nearly all of the known sensitive mixed conifer forest 

ecosystems receive total N deposition levels above 3.1 N kg/ha/yr, which is the ecological benchmark for 

changes in lichen species. Lichens are sentinels for broader ecosystem change in terrestrial systems. 

Some portions of the Sierra Nevadas receive total N deposition levels above 5.2 N kg/ha/yr, which is the 

ecological benchmark for shifts in the dominant species of lichen from acidophytic to tolerant species. In 

addition, in Coastal Scrub Sage ecosystems in California, N deposition exceeds the 3.3 N kg/ha/yr 

benchmark above which nitrogen is no longer a limiting nutrient, leading to potential alterations in 

ecosystem composition. Because excessive N deposition and effects are observed in areas where, under 

recent conditions, NOx ambient concentrations are at or below the current NOx secondary standards, we 

initially conclude those standards are not adequate to protect against anticipated adverse impacts from N 

nutrient enrichment in sensitive ecosystems (systems where N is limiting) that are not managed for 

commercial agricultural and forest production. 

Sulfur deposition is also linked with the formation of methylmercury. The production of methylmercury 

in aquatic ecosystems requires sulfate as well as mercury. The evidence is sufficient to infer a causal 

relationship between sulfur deposition and increased mercury methylation in wetlands and aquatic 

environments. However, while the production of methylmercury requires the presence of sulfate and 

mercury, the amount of methylmercury produced varies with oxygen content, temperature, pH, and 

supply of labile organic carbon. Due to limits in data, we are unable to assess the adequacy of the 

existing standards in protecting against effects associated with increased mercury methylation. 

Conceptual Design of an Ecologically Relevant Standard 

The overall concept for ecologically relevant standards recognizes that the fundamental welfare effects 

associated with ambient NOx and SOx occur through the process of deposition to sensitive ecosystems. 

There are four main components to the conceptual design of the standard: atmospheric and ecological 

indicators, deposition metrics, functions that relate indicators to deposition metrics and factors that 

modify the functions. In this policy assessment, the focus is on developing the conceptual design for a 

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Draft Policy Assessment for the Review of the Secondary NAAQS for NOx and SOx: 

Executive Summary 

standard that protects against effects associated with acidifying deposition of NOx and SOx in aquatic 

ecosystems, but this general conceptual framework is intended to apply to a broader set of potential 

endpoints. 

For the conceptual design of an aquatic acidification standard, ANC is suggested as the ecological 

indicator. ANC is suggested as the ecological indicator because it is the most widely used chemical 

indicator of acid sensitivity in aquatic ecosystems and has been found through numerous studies to be the 

best single indicator of the biological response and health of aquatic communities in acid-sensitive 

systems. Furthermore, ANC can be directly linked to both underlying water chemistry, e.g. pH and 

aluminum, and to biological impairment, specifically the number of fish species in a water body. 

Acidification models represent the ecological response relationship between ANC and deposition of N 

and S. Acidification models are designed for the catchment scale. However, for consideration of a 

national standard, aggregation to a broader spatial scale is desirable; therefore a method to evaluate 

critical loads (the levels of deposition of N and S below which defined levels of harmful effects on 

specified sensitive elements of the environment do not occur) across the national landscape is presented. 

The atmospheric transformation functions then convert deposition to ambient concentrations of NOx and 

SOx. 

Acidification models relate ANC to deposition of N and S at the catchment scale. We suggest using an 

acidification model that incorporates environmental variables that modify the ecological response 

relationship. This includes a variable to account for nitrogen uptake by ecosystems. The acidification 

model can be used to calculate critical loads for individual catchments based on a selected level of ANC. 

The load of deposition that causes a selected level of ANC varies across the nation depending on the 

characteristics of the catchment, such as base cation weathering rates, nitrogen retention and the level of 

naturally occurring organic acids. 

Although critical loads for a selected level of ANC will vary catchment by catchment it is not practical for 

a national standard to evaluate every catchment in the U.S. Therefore, we propose two general approaches 

to establish critical loads. One approach is to develop a national distribution of critical loads over all 

levels of sensitivity, recognizing that there is a high degree of heterogeneity in acid sensitivity even at 

relatively small spatial scales. The second approach is to subdived the landscape of the U.S.into acidsensitivity categories, such that within a category there are generally similar acid sensitivity 

characteristics. Each national acid-sensitivity category is represented by a population of catchments for 

which critical loads at a specified ANC limit are calculated. This second draft policy assessment explores 

a number of methods for developing the acid sensitivity categories. 

National acid-sensitivity categories should be based on features that govern ecological sensitivity to 

acidification. Areas that have similar underlying geology, mineral weathering rates, and hydrology 

should show similar sensitivity to NOx and SOx deposition. Ecoregions are useful geographic defiitions 

that holistically incorporate a number of important factors related to acid sensitivity, including geology, 

physiography, vegetation, climate, soils, land use, and hydrology. As such, the determination of acidsensitivity categories begins with aggregation of lakes and streams by ecoregion. From there, acid 

sensitivity is further characterized using measured ANC. As noted earlier, ANC is a good indicator of the 

overall sensitivity of a water body to acidification. ANC measurement is also widely available in the 

U.S., making it a useful metric for further classifying ecoregions. The policy assessment explores several 

methods for using ANC to classify ecoregions using both simple and more complex methods. 

Once acid-sensitivity categories are defined, a sample of catchments will be selected to represent each 

category. The acidification models will be used to evaluate the critical load for each catchment in the 

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Draft Policy Assessment for the Review of the Secondary NAAQS for NOx and SOx: 

Executive Summary 

population at the selected level of ANC. There is a distribution of critical load values within each acidsensitivity category reflecting the range of sensitivity of catchments within each category. The goal of 

aggregating critical loads from multiple catchments is to develop an appropriately representative 

deposition value based on the distribution of critical loads, called the deposition metric, which protects a 

percentage of the population of water bodies within a national acid-sensitivity category from exceeding 

their critical load for the target value of ANC. 

Distributions of catchment level critical loads are based on a combination of previously conducted steadystate critical load modeling and new critical load modeling conducted as part of this policy assessment. 

To ensure the population of water bodies included in the analysis were those sensitive to acidity caused 

by atmospheric deposition, several criteria were applied to the critical loads dataset to remove catchments 

in which organic acids, acid mine drainage or naturally low base cation weathering caused acidification. 

Once a deposition metric is calculated, the value is modified by addition of a term to represent the amount 

of N that will be taken up by vegetation, immobilized in soil or degassed from the ecosystem. Next, a 

tradeoff curve for N and S deposition is generated for the acid sensitivity category. This function 

(illustrated below) is characterized by three nodes: 1) the maximum of amount of N deposition when S 

deposition equals zero 2) the amount of N deposition that will be captured by the ecosystem before it 

leaches and 3) the maximum amount of S sulfur deposition considering the N captured by the ecosystem. 

The function represents all pairs of N and S deposition that will equal the deposition metric for acidifying 

deposition for a specific target ANC. 

The depositional load function

Reduced forms of nitrogen, such as ammonia, are either taken up by plants and microbes or converted to 

nitrate in the environment and use up the assimilative capacity of ANC at the same rate as oxidized forms 

of nitrogen deposition; therefore, deposition of reduced nitrogen must be accounted for in the watershed. 

The suggested approach is to subtract the loadings of reduced forms of nitrogen derived for a given 

spatial area from the deposition metric that represents a selected percentage of critical loads for a given 

population, such that the resultant deposition metric is for sulfur and oxidized nitrogen only. The policy 

assessment explores several methods for specifying the loadings of reduced nitrogen for specific 

geographic areas. 

Deposition is related to ambient concentrations of NOx and SOx through deposition velocity, which is the 

rate at which an ambient pollutant is deposited. Deposition velocity varies over time and location, and is 

affected by land use conditions and meteorology. Our conceptual model requires conversion of 

deposition of N and S into ambient concentrations of NOx and SOx. Since the policy objective is to set an 

ambient air quality standard for total oxidized sulfur and nitrogen, and this is also the chemical resolution 

provided by the ecosystem models, it is convenient to use conversion factors based on available estimates 

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Draft Policy Assessment for the Review of the Secondary NAAQS for NOx and SOx: 

Executive Summary 

of concurrent deposition and ambient concentrations. The ratio of total deposition of NOx and SOx to 

ambient concentrations of NOx and SOx, respectively are referred to as deposition transference ratios. 

Applying the estimated deposition transference ratios to depositional load tradeoff curves leads to parallel 

tradeoff curves for ambient NOx and SOx (see figure below). 

The national ambient air quality standard must be able to tie ANC to deposition and deposition to ambient 

air concentrations, incorporating ecological conditions and the contribution of reduced nitrogen. To 

incorporate all of these aspects, we develop an index that will provide a consistent standard nationally that 

is directly expressed in terms of concentrations of NOx and SOx. The form of this standard is referred to 

as the Atmospheric Acidification Protection Index (AAPI), which can be applied across the nation to 

convey the protection of aquatic ecosystems from acidification due to atmospheric deposition. 

The AAPI represents the level of protection against the effects of acidification given local ecological 

conditions, the level of reduced nitrogen being deposited, and the levels of NOx and SOx that are limited 

by the standard. The AAPI is linked to a target ANC for a chosen percent of lakes and streams, and 

determines the combinations of NOx and SOx that will jointly result in the target ANC, taking into 

account uncertainties and other factors. 

This AAPI can also be expressed as the tradeoff curves which show the combinations of NOx and SOx

that meet the standard, generated for specific values of AAPI, and provide a representation of the 

standards in terms of atmospheric concentrations. These tradeoff curves will vary across the U.S. based 

on ecosystem sensitivity, reduced nitrogen deposition levels, and other factors. An example tradeoff 

curve for two different percentiles of protection at a target ANC of 50 μeq/L is provided below. 

0

50

100

0 50 100

N (meq/m2/yr)

SO2+SO4 (meq/m2/yr)

90%

75%

Neco

NHx dep = 4

NHx dep = 54

0

50

100

0 50 100

N (meq/m2

/yr)

SO2+SO4 (meq/m2/yr)

90%

75%

Neco

0

50

100

0 50 100 150

NOy (meq/m2/yr)

SO2+SO4 (meq/m2/yr)

90% and NHx=5

90% and NHx=54

75% & NHx =5

75% & NHx =54

0

1

2

3

4

5

012345

NOy (ug/m3)

SO2+SO4(ug/m3)

90% and NHx=5

90% and NHx=54

75% & NHx =5

75% & NHx =54

Adirondack air quality

A

B

C

D

Graphs demonstrate the steps to develop a tradeoff curve for the AAPI including A) N vs. 

SO2+SO4 from the deposition metric and modified with Neco B) Two loads of NHx indicated C) 

after subtracting NHx load the resultant NOy vs SO2+SO4 curve, D) After applying the 

transference ratio the resultant air concentrations.

0

50

100

0 50 100

N (meq/m2/yr)

SO2+SO4 (meq/m2/yr)

90%

75%

Neco

NHx dep = 4

NHx dep = 54

0

50

100

0 50 100

N (meq/m2

/yr)

SO2+SO4 (meq/m2/yr)

90%

75%

Neco

0

50

100

0 50 100 150

NOy (meq/m2/yr)

SO2+SO4 (meq/m2/yr)

90% and NHx=5

90% and NHx=54

75% & NHx =5

75% & NHx =54

0

1

2

3

4

5

012345

NOy (ug/m3)

SO2+SO4(ug/m3)

90% and NHx=5

90% and NHx=54

75% & NHx =5

75% & NHx =54

Adirondack air quality

A

B

C

D

Graphs demonstrate the steps to develop a tradeoff curve for the AAPI including A) N vs. 

SO2+SO4 from the deposition metric and modified with Neco B) Two loads of NHx indicated C) 

after subtracting NHx load the resultant NOy vs SO2+SO4 curve, D) After applying the 

transference ratio the resultant air concentrations. 

Options for Elements of the Standards 

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Draft Policy Assessment for the Review of the Secondary NAAQS for NOx and SOx: 

Executive Summary 

Ambient air quality indicators other than NO2 and SO2 should be considered as the appropriate pollutant 

indicators for protection against the acidification effects associated with deposition of NOx and SOx. This 

consideration is based on the recognition that all forms of oxidized nitrogen and sulfur in the atmosphere 

contribute to deposition and resulting acidification, and as such concentrations of NO2 and SO2 are 

incomplete indicators. Furthermore, concentration of NOy is proposed as an appropriate indicator for 

oxides of nitrogen. The sum of concentrations of SO2 and SO4 is proposed as an appropriate indicator for 

oxides of sulfur. 

Welfare effects associated with acidification result from annual cumulative deposition of nitrogen and 

sulfur, reflected in effects on the chronic ANC level (measured as annual or multiyear average ANC). 

Short-term (i.e., hours or days) episodic changes in water chemistry can also have significant biological 

effects. Episodic chemistry refers to conditions during precipitation or snowmelt events. Surface water 

chemistry has lower pH and acid neutralizing capacity (ANC) during these events than under baseflow 

conditions. One of the most important effects of acidifying deposition on surface water chemistry is the 

short-term change in chemistry that is termed “episodic acidification.” While ecosystems are also 

affected by episodic increases in acidity due to pulses of acidity during high rainfall periods and 

snowmelts, protection against these episodic acidity events can be achieved by establishing a higher 

chronic ANC level. Episodic acidification can result from either shorter term deposition episodes, or 

from longer term deposition on snowpack. Snowmelt can release stored N deposited throughout the 

winter, leading to episodic acidification in the absence of increased deposition during the actual episodic 

acidification event. Long term (3 to 5 year average) ambient NOx and SOx concentrations are appropriate 

to provide protection against low chronic ANC levels and episodic acidification. 

The current forms of the secondary standards for NOx and SOx do not take into account the combined 

contributions of NOx and SOx in the causation of effects associated with acidification of aquatic 

ecosystems. Based on the causal linkages between NOx and SOx, deposition of N and S, and the indicator 

of acidification, ANC, the current forms should be replaced with an atmospheric acidification protection 

index (AAPI), which reflects the important roles of underlying ecosystem characteristics, determinants of 

deposition, and deposition of reduced nitrogen in determining the potential effects from deposition of 

NOx and SOx. 

The value of AAPI can be calculated for any observed values of NOx and SOx. However, the level of the 

standard for AAPI should reflect a wide number of factors, including desired level of protection indicated 

by a target ANC limit, the target percentile of water bodies to achieve the target ANC, and the various 

factors and uncertainties involved in specifying all of the other aspects of the standard, such as the acid 

sensitivity classification method, the specification of deposition of reduced nitrogen, nitrogen retention, 

the deposition transference ratios, and the averaging time. The administrator may choose an AAPI level 

reflecting an ANC level higher or lower than the target ANC limit to account for the combined effect of 

all of the components of the standard and their related uncertainties. The resulting AAPI, in the context of 

the overall standard, will reflect her informed judgment as to a standard that is sufficient but not more 

than necessary to protect against adverse public welfare effects. 

Within the AAPI form, EPA will specify the parameter values for all elements excepting NOx and SOx, 

which will be the measured atmospheric indicators. The values for pre-industrial base cation weathering, 

nitrogen retention and uptake, and runoff, are based on the same inputs used to develop the deposition 

metrics. The values for reduced nitrogen and the deposition transference ratios will be calculated using 

output from the most up-to-date version of EPA’s Community Multiscale Air Quality (CMAQ) model. 

EPA is considering methods to account for the dynamic nature of deposition of reduced nitrogen in 

specifying the NOx and SOx standards. 

ES-9
Draft Policy Assessment for the Review of the Secondary NAAQS for NOx and SOx: 

Executive Summary 

As demonstrated by the tradeoff curves, multiple combinations of concentrations of NOx and SOx can 

yield the same value of the AAPI. No single combination of NOx and SOx will produce a particular value 

of AAPI in all locations. Measured concentrations of annual average NOx and SOx necessary to meet the 

standards are thus expressed conditionally by the AAPI and not by fixed quantities. 

A target ANC limit based on a desired level of protection is an important input to the decisions of the 

level of AAPI and the percent of ecosystems to be protected. Specific levels of ANC are associated with 

differing levels of ecosystem impairment, with higher levels of ANC resulting in fewer ecosystem 

impacts, and lower levels resulting in both higher intensity of impacts and a broader set of impacts. For 

example, the number of fish species present in a waterbody 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. This relationship is illustrated in the following figure. 

Number of fish species per lake or stream versus ANC level and 

aquatic status category for lakes in the Adirondack Case Study 

Area

The target ANC level specified in designing the standard is only one part in determining the overall 

protectiveness of the standard. The degree of protectiveness is based on all elements of the standard, 

including the target ANC selected by the Administrator, the size of the spatial areas over which the 

standard is applied, the percent of aquatic ecosystems targeted within a spatial area that is selected by the 

Administrator to achieve the selected ANC level, and the underlying parameters of the AAPI, including 

the atmospheric indicator, the critical load models used to determine pre-industrial base cation levels, the 

calculated values for the deposition transference ratios, and the calculated value for deposition of reduced 

nitrogen. There are widely varying degrees of uncertainty associated with all of these elements, some 

being much more certain and others being much less certain. The specified target ANC level is a crucial 

part of developing a standard that is requisite to protect public welfare, but it is the overall design and 

content of the standard that must be considered in judging the adequacy of protection it provides. 

The secondary NAAQS will reflect the public welfare policy judgments of the Administrator, based on 

the science, as to the level of air quality which is requisite to protect the public welfare from any known 

or anticipated adverse effects associated with the pollutant in the ambient air. In certain naturally acidic 

ecosystems, even though the ecological benchmarks are exceeded, e.g. ANC may be quite low; NOx and 

ES-10
Draft Policy Assessment for the Review of the Secondary NAAQS for NOx and SOx: 

Executive Summary 

SOx are not contributing to effects because those systems have chronic natural acidity and will not benefit 

from reductions in atmospheric deposition. The secondary NAAQS are not intended to provide 

protection in these types of naturally acidic systems. Instead, the secondary NAAQS are focused on 

providing protection in areas where ambient NOx and SOx are resulting in effects in ecosystems with low 

natural levels of acidification that are highly sensitive to additional inputs of acid deposition. The 

approaches for specifying populations of critical loads to develop deposition metrics explicitly excludes 

lakes and streams that are naturally acidic and those not likely to benefit from decreases in atmospheric 

deposition. 

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. Overall, there is little uncertainty that significant 

effects on aquatic biota are occurring at ANC levels below 20 μeq/L. Based on the field data from the 

Adirondacks and Shenendoah case study areas, ANC levels less than 50 μeq/L are adverse to ecosystem 

health, and are likely to lead to reductions in ecosystem services related to recreational fishing. However, 

the types of effects, specific species, and prevalence of effects across water bodies in the U.S. is more 

uncertain at ANC levels between 20 and 50 μeq/L. Targeting ANC levels between 50 and 100 μeq/L 

would provide additional protection; however, uncertainties regarding the additional reduction in adverse 

welfare effects are much larger for target ANC levels above 50 μeq/L. 

Specifying an appropriate range of levels for an AAPI standard that is designed and specified as discussed 

above involves consideration of the degree to which any specific AAPI would lead to achieving the 

desired ANC level, and a judgment as to the degree of protection of public welfare that is warranted. 

Selection of a range of AAPI and selection of a specific level of AAPI within that range should 

incorporate a wide number of considerations, including the percent of water bodies within acid sensitive 

areas that the Administrator determines should be protected at the targeted ANC level. 

In determining the requisite level of protection for the public welfare from effects on aquatic ecosystems, 

the Administrator will need to weigh the importance of the predicted risks of these effects in the overall 

context of public welfare protection, along with a determination as to the appropriate weight to place on 

the associated uncertainties and limitations of this information. 

Co-Protection for Other Effects Provided by an Acidification Standard 

To understand the level of protection provided by a NOx/SOx standard based on aquatic acidification to 

protect against terrestrial acidification effects, we compared the critical loads for lakes and streams that 

would maintain an aquatic ANC of 50 to the critical loads to maintain either a terrestrial Bc:Al ratio of 1.2 

or 10 averaged across a watershed area. 

Results for the Adirondacks showed that critical loads for 29 lakes at an ANC of 50 were lower for 13 of 

those lakes than the critical load for the terrestrial watershed areas at a Bc:Al ratio of 10 and for 21 lakes 

at a Bc:Al ratio of 1.2. Perhaps more significant was the result that 13 of the 16 lakes in the highly and 

moderately sensitive areas had a lower critical load than the Bc:Al 10 areas and 16 of 16 lakes in the 

highly and moderately sensitive areas had lower critical loads than the Bc:Al 1.2 areas. The Shenandoah 

region reflected similar results. 

In general, the aquatic critical acid loads offered greater protection to the watersheds than did the 

terrestrial critical loads. Generally in situations where the terrestrial critical loads were more protective, 

the lakes or streams in the watershed were rated as having “Low Sensitivity” or “Not Sensitive” to 

acidifying nitrogen and sulfur deposition. Conversely, when the water bodies were more sensitive to 

ES-11
Draft Policy Assessment for the Review of the Secondary NAAQS for NOx and SOx: 

Executive Summary 

deposition (“Highly Sensitive” or “Moderately Sensitive”), the aquatic critical acid loads generally 

provided a greater level of protection against acidifying nitrogen and sulfur deposition in the watershed. 

Initial Conclusions 

In this current review, important new information has become available since the last reviews (1996 for 

NOx, and 1988 for SOx) that supports revising the current NOx and SOx standards. Specifically, the ISA 

has concluded that there are causal relationships between NOx and SOx acidifying deposition and effects 

on aquatic and terrestrial ecosystems, and the ISA and REA provide substantial quantitative evidence of 

effects occurring in locations that meet the current NO2 and SO2 standards. In addition, observational 

data and rigorous atmospheric modeling has become available regarding the role of both nitrogen and 

sulfur deposition in acidification of sensitive water bodies. This information is sufficient to inform the 

development of revised secondary standards for NOx and SOx to protect against the effects of aquatic 

acidification in sensitive ecosystems. Additional information is needed to set separate standards to 

protect terrestrial ecosystems from acidification effects. While there is also new information available on 

the role of nitrogen deposition on nutrient enrichment effects in sensitive terrestrial and aquatic 

ecosystems, and the ISA concludes there is a causal relationship between NOx and nutrient enrichment 

effects, for this second draft policy assessment, we have focused on acidification effects due to the 

substantially greater amount of information available to inform the development of secondary standards. 

We highlight the progress made in considering the joint nature of ecosystem responses to acidifying 

deposition of NOx and SOx, and note that the ability to consider revisions to the NOx and SOx secondary 

standards has been enhanced by our ability to consider a joint standard for NOx and SOx to protect against 

acidification effects. The development of an appropriate form of the standard linked to a common 

indicator of aquatic acidification, ANC, is also a significant step forward, as it allows for development of 

a standard for aquatic acidification designed to provide generally the same degree of protection across the 

country, while still reflecting the underlying variability in ecosystem sensitivity to acidifying NOx and 

SOx deposition. 

We provide the following initial conclusions regarding the NOx and SOx secondary standards: 

 The available effects-based evidence for aquatic and terrestrial acidification and nutrient 

enrichment suggests consideration of NOx and SOx standards that are at least as protective as the 

current standard. Consideration of joint standards for NOx and SOx is appropriate given the 

common atmospheric processes governing the deposition of NOx and SOx to sensitive 

ecosystems. 

 On the basis of the acidification and nutrient enrichment effects that have been observed to still 

occur under current ambient conditions and those predicted to occur under the scenario of just 

meeting the current secondary NAAQS, we find support for consideration that the current 

secondary NAAQS are inadequate to protect the public welfare from known and anticipated 

adverse welfare effects from aquatic and terrestrial acidification associated with deposition of 

NOx and SOx. 

 We find support for consideration that current levels of NOx and SOx are associated with 

deposition that leads to ANC values below benchmark values that cause ecological harm and 

losses in ecosystem services in sensitive ecosystems, including significant mortality in sensitive 

aquatic biota and losses in fish species richness, which is associated with reductions in 

recreational fishing services, among others. 

 We find support for consideration that current levels of ambient NOx and SOx are associated with 

deposition that leads to Bc:Al values below benchmark values for terrestrial acidification that 

cause ecological harm and losses in ecosystem services in sensitive ecosystems, including losses 

ES-12
Draft Policy Assessment for the Review of the Secondary NAAQS for NOx and SOx: 

Executive Summary 

ES-13

in tree health and growth, which are associated with reductions in timber production, among other 

services. 

 We suggest that effects due to aquatic acidification are most suitable for defining secondary 

standards for NOx and SOx. We note that in developing a standard designed to protect against the 

effects of aquatic acidification due to deposition of NOx and SOx, the resulting standards may not 

provide adequate protection against known effects associated with nutrient enrichment in 

sensitive aquatic and terrestrial ecosystems or acidification in sensitive terrestrial ecosystems. 

 Based on the causal linkages between NOx and SOx, deposition of N and S, and the indicator of 

acidification, ANC, consideration should be given to an additional secondary standard with a 

form defined by an atmospheric acidification protection index (AAPI), which reflects the 

important roles of underlying ecosystem characteristics, determinants of deposition, and 

deposition of reduced nitrogen in determining the potential effects from deposition of NOx and 

SOx. 

 Staff has concluded, based on the evidence and risk based information, and consideration of 

information related to definitions of adversity, that: 

 a target level of ANC of 20 μeq/L will protect against significant losses in fish mortality in 

many sensitive lakes, but will place less weight on protection against losses in aquatic 

biodiversity, and will be less protective against potential acidification episodes, 

 a target level of ANC of 50 μeq/L will protect against significant mortality in aquatic 

organisms and loss of fish health and biodiversity in sensitive lakes and streams, and will 

give weight to considerations of uncertainties in the time to recovery of aquatic ecosystems, 

 target levels of ANC above 50 μeq/L may provide additional protection against declines in 

fitness of sensitive species (e.g., brook trout, zooplankton), however, overall health of aquatic 

communities may not be impacted. 
 TABLE OF CONTENTS 

List of Figures.................................................................................................................. vii 

 List of Tables .................................................................................................................... xi 

 List of Acronyms/Abbreviations................................................................................... xiii 

 List of Key Terms........................................................................................................... xvi 

1 INTRODUCTION.......................................................................................................... 1-1 

 1.1 DEFINITIONS OF NOx AND SOx FOR THIS ASSESSMENT........................... 1-3 

 1.2 POLICY OBJECTIVES........................................................................................... 1-8 

 1.3 CRITICAL POLICY ELEMENTS........................................................................ 1-10 

 1.4 HISTORICAL CONTEXT .................................................................................... 1-12 

 1.4.1 History of NOx and SOx NAAQS Review................................................ 1-12 

 1.4.2 History of Related Assessments and Agency Actions............................... 1-14 

 1.5 PROPOSED CONCEPTUAL FRAMEWORK FOR COMBINED 

 NOX SOX STANDARDS...................................................................................... 1-17 

 1.6 POLICY RELEVANT QUESTIONS.................................................................... 1-21 

2 KNOWN OR ANTICIPATED ADVERSE ECOLOGICAL EFFECTS.................. 2-1 

2.1 Acidification: Evidence of effects on structure and function of terrestrial and 

freshwater ecosystems ............................................................................................. 2-2 

 2.1.1 What is the nature of acidification related ecosystem responses to 

 reactive nitrogen and sulfur deposition? ...................................................... 2-3 

 2.1.2 What types of ecosystems are sensitive to such effects? 

 In which ways are these responses affected by atmospheric, ecological, and 

landscape factors? ........................................................................................ 2-6 

 2.1.3 What is the magnitude of ecosystem responses to acidifying deposition? .. 2-6 

 2.1.4 What are the key uncertainties associated with acidification?................... 2-17 

 2.2 Nitrogen-enrichment: Evidence of effects on structure and function of terrestrial and 

 freshwater ecosystems ........................................................................................... 2-19 

 2.2.1 What is the nature of terrestrial and freshwater ecosystem responses to 

reactive nitrogen and sulfur deposition? .................................................... 2-20 

i
 2.2.2 What types of ecosystems are sensitive to such effects? How are these 

responses affected by atmospheric, ecological, and landscape factors...... 2-22 

 2.2.3 What is the magnitude of ecosystem responses to nitrogen deposition? ... 2-23 

 2.2.4 What are the key uncertainties associated with nutrient enrichment? ....... 2-32 

 2.3 What ecological effects are associated with gas-phase NOx and SOx?.................. 2-33 

 2.3.1 What is the nature of ecosystem responses to gas-phase 

 nitrogen and sulfur? ................................................................................... 2-33 

 2.3.2 What types of ecosystems are sensitive to such effects? How are these 

responses affected by atmospheric, ecological, and landscape factors?.... 2-34 

 2.3.3 What is the magnitude of ecosystem responses to gas phase effects 

 of NOx and SOx?....................................................................................... 2-34 

 2.4 SUMMARY........................................................................................................... 2-35 

 2.5 REFERENCES ...................................................................................................... 2-37 

3 CONSIDERATIONS OF ADVERSITY TO PUBLIC WELFARE.......................... 3-1 

3.1 How do we characterize adversity? What are the relevant definitions and how are 

they addressed in this document?............................................................................. 3-1 

 3.1.1 What are the benchmarks for adversity from other sources?....................... 3-2 

 3.1.2 Other EPA Programs and Federal Agencies................................................ 3-4 

 3.2 What are ecosystem services and how does this concept relate to 

 public welfare?....................................................................................................... 3-10 

 3.3 Applying Economic Valuation to Ecosystem Services.......................................... 3-17 

 3.3.1 Ecosystem services and links to public welfare..........................................3.17 

 3.3.2 Economics as a framework to illustrate changes in public welfare ........... 3-17 

 3.3.3 The role of economics in defining “adversity” .......................................... 3-19 

 3.4 Effects of Acidification and Nutrient Enrichment on Ecosystem Services ........... 3-24

 3.4.1 Evidence for adversity related to aquatic acidification.............................. 3-26 

 3.4.2 Evidence for Adversity Related to Terrestrial Acidification ..................... 3-31 

 3.4.3 Evidence for Adversity Related to Aquatic Nutrient Enrichment ............. 3-35 

 3.4.4 Evidence for Adversity Related to Terrestrial Nutrient Enrichment ......... 3-39 

 3.5 REFERENCES ...................................................................................................... 3-42 

4 ADDRESSING THE ADEQUACY OF THE CURRENT STANDARDS ............... 4-1 

ii
4.1 Are the structures of the current NOX and SOX secondary standards based on 

 relevant ecological indicators such that they are adequate to determine and protect 

 public welfare against adverse effects on ecosystems? ........................................... 4-1 

 4.2 To what extent are the structures of the current NOX and SOX secondary standards 

 meaningfully related to relevant ecological indicators of public welfare effects? .. 4-3 

 4.3 To what extent do current monitoring networks provide a sufficient basis for 

 determining the adequacy of current secondary NOX and SOX standards? ............. 4-6 

 4.3.1 To what extent does the NADP monitoring network provide an adequate 

characterization of deposition and what are the major limitations?........... 4-13 

 4.3.2 What are the relative strengths and important gaps in existing atmospheric 

monitoring networks to address a combined NOX/SOX 

 secondary standard? ................................................................................... 4-15 

 4.3.3 How do we characterize deposition through monitoring and models?...... 4-16 

 4.4 What is our best characterization of atmospheric concentrations of NOY and SOX, 

 and deposition of N and S? .................................................................................... 4-18 

 4.4.1 What are the current atmospheric concentrations of reactive nitrogen, NOy, 

reduced nitrogen, NHX, sulfur dioxide, SO2, and sulfate, SO4? ................ 4-18 

 4.5 Are adverse effects on the public welfare occurring under current air quality 

conditions for NO2 and SO2 and would they occur if the nation met the current 

standards?............................................................................................................... 4-38 

 4.5.1 To what extent do the current NOX and SOX secondary standards provide 

protection from effects associated with deposition of atmospheric NOX and 

SOX which results in acidification in sensitive aquatic and terrestrial 

ecosystems?................................................................................................ 4-40 

 4.5.2 To what extent does the current NOX secondary standard provide protection 

from effects associated with deposition of oxidized nitrogen from 

atmospheric NOX, which results in nutrient enrichment effects in sensitive 

aquatic and terrestrial ecosystems?............................................................ 4-45 

 4.5.3 Aquatic Nutrient Enrichment..................................................................... 4-47 

 4.5.4 Terrestrial Nutrient Enrichment................................................................. 4-48 

iii
4.6 To what extent do the current NOX and/or SOX secondary standards provide 

protection from other ecological effects (eg. mercury methylation) associated with 

the deposition of atmospheric NOX, and/or SOX? ................................................. 4-49 

4.7 REFERENCES ...................................................................................................... 4-51 

5 OPTIONS FOR ELEMENTS OF THE STANDARD................................................ 5-1 

 5.1 What atmospheric indicators of oxidized nitrogen and sulfur are appropriate for use 

in a secondary NAAQS that provides protection for public welfare from exposure 

related to deposition of NOx and SOx? ................................................................... 5-3 

 5.2 What is the appropriate averaging time for the air quality indicators NOy and SOx to 

provide protection of public welfare from adverse effects from 

 aquatic acidification? ............................................................................................... 5-4 

 5.3 What form(s) of the standard are most appropriate to provide protection of sensitive 

ecosystems from the effects of acidifying deposition related to ambient NOx and 

SOx concentrations? ................................................................................................ 5-5 

 5.3.1 Conceptual Design of the Form: General Overview.................................... 5-5 

 5.3.2 Conceptual design of the form: Linking the ecosystem indicator 

 to deposition............................................................................................... 5-11 

 5.3.3 Conceptual Design: Linking Deposition to Atmospheric Concentration . 5-53 

 5.3.4 AAPI .......................................................................................................... 5-61 

 5.4 Options for specifying the targets for the ecological indicator for aquatic 

acidification (ANC)................................................................................................ 5-68 

 5.4.1 What levels of impairment are related to alternative target levels of 

 ANC? ......................................................................................................... 5-69 

 5.4.2 Additional information on Target ANC levels .......................................... 5-84 

 5.4.3 Adversity of ecological impacts associated with alternative target ANC 

levels .................................................................................................................. 5-85 

 5.5 Options for specifying the targets for the deposition metric.................................. 5-87 

 5.6 REFERENCES ...................................................................................................... 5-90 

6 CO-PROTECTION FOR OTHER EFFECTS USING STANDARDS TO 

PROTECT AGAINST ACIDIFICATION .................................................................. 6-1 

iv
6.1 To what extent would a standard specifically defined to protect against aquatic 

acidification likely provide protection from terrestrial acidification? ..................... 6-1 

 6.2 To what extent would a standard specifically defined to protect against aquatic 

acidification likely provide protection from terrestrial nutrient enrichment?.......... 6-4 

6.3 To what extent would a standard specifically defined to protect against aquatic 

acidification likely provide protection from aquatic nutrient enrichment?.............. 6-6 

 6.4 REFERENCES ........................................................................................................ 6-7 

7 Evaluation of Uncertainty and Variability in the Context of an AAPI standard, 

including Model Evaluation, Sensitivity Analyses, and Assessment of Information 

Gaps ...................................................................................................................... 7-1

7.1 INTRODUCTION and PURPPOSE........................................................................ 7-1 

 7.2 Uncertainty associated with ecosystem effects and dose – response relationships . 7-5 

 7.3 Uncertainty in benefits estimates ............................................................................. 7-7 

 7.4 CMAQ application and evaluation .......................................................................... 7-9 

 7.4.1 Overview of cmaq model application .......................................................... 7-9 

 7.4.2 CMAQ Evaluation, Sensitivity and Variability Analyses ......................... 7-10 

 7.4.3 Variability and sensitivity of CMAQ generated component ..................... 7-12 

 7.5 Sensitivity of AAPI to component parameters ...................................................... 7-30 

 7.6 Uncertainty in Critical Load and ANC modeling .................................................. 7-30 

 7.6.1 MAGIC modeling ...................................................................................... 7-30 

 7.6.2 SSWC modeling......................................................................................... 7-31 

 7.7 Modeling and Data Gaps ....................................................................................... 7-34 

 7.8 Summary and Conclusions .................................................................................... 7-36 

 7.9 REFERENCES ...................................................................................................... 7-40 

8 AMBIENT AIR MONITORING.................................................................................. 8-1 

8.1 What are the appropriate ambient air indicators to consider in developing the 

standards?................................................................................................................. 8-1 

 8.2 Reactive Oxidized Nitrogen and Sulfur Species...................................................... 8-5 

 8.3 What measurements would be used to characterize NOy and SOx ambient air 

concentrations for the purposes of the AAPI based standard?................................. 8-7 

 8.4 What additional complementary measurements are recommended?....................... 8-8 

v
8.5 What sampling frequency would be required? ........................................................ 8-9 

 8.6 What are the spatial scale issues associated with monitoring for compliance, and 

how should these be addressed? ............................................................................ 8-10 

 8.7 What specific monitoring methods would be used? .............................................. 8-10 

 8.8 REFERENCES ...................................................................................................... 8-14 

9 INITIAL CONCLUSIONS ........................................................................................... 9-1 

9.1 CONSIDERATION OF ALTERNATIVE STANDARDS ..................................... 9-3 

 9.1.1 Indicators...................................................................................................... 9-5 

 9.1.2 Averaging Times.......................................................................................... 9-5 

 9.1.3 Form............................................................................................................. 9-6 

 9.1.4 Alternative levels of the AAPI standard ...................................................... 9-9 

 9.1.5 Additional protections for ecosystems against the effects of terrestrial 

acidification and terrestrial and aquatic nutrient enrichment..................... 9-15 

 9.1.6 Summary of options................................................................................... 9-16 

 9.2 CONCLUSIONS.................................................................................................... 9-16 

 9.3 REFERENCES ...................................................................................................... 9-21 

APPENDICES 

 Appendix A Additional Information for Chapter 5 

 Appendix B Analysis of Critical Loads, Comparing Aquatic and Terrestrial 

 Acidification 

 Appendix C Elasticity and ANOVA AAPI sensitivity analyses 

vi
List of Figures 

Figure 1-1 Framework of an alternative secondary standard .............................................. 1-21 

Figure 2-1 Conceptual model of direct and indirect acidification effects on aquatic biota... 2-4 

Figure 2-2 Average NO3

- concentrations (orange), SO42- concentrations (red), and ANC 

(blue) across the 44 lakes in the Adirondack Case Study Area modeled using 

MAGIC for the period 1850 to 2050 ................................................................. 2-11 

Figure 2-3 ANC concentrations of preacidification (1860) and current (2006) conditions 

based on hindcasts of 44 lakes in the Adirondack Case Study Area modeled using 

MAGIC .............................................................................................................. 2-12 

Figure 2-4 Critical loads of acidifying deposition that each surface water location can receive 

in the Adirondack Case Study Area while maintaining or exceeding an ANC 

concentration of 50 μeq/L based on 2002 data.. ................................................ 2-13 

Figure 2-5 Average NO3

- concentrations orange), SO42-concentrations (red), and ANC (blue) 

levels for the 60 streams in the Shenandoah Case Study Area modeled using 

MAGIC for the period 1850 to 2050 ................................................................. 2-15 

Figure 2-6 ANC levels of 1860 (preacidification) and 2006 (current) conditions based on 

hindcasts of 60 streams in the Shenandoah Case Study Area modeled using 

MAGIC .............................................................................................................. 2-15 

Figure 2-7 Critical loads of surface water acidity for an ANC of 50 μeq/L for Shenandoah 

Case Study Area streams. Each dot represents an estimated amount of acidifying 

deposition (i.e., critical load) that each stream’s watershed can receive and still 

maintain a surface water ANC >50 μeq/L. ........................................................ 2-16 

Figure 2-8 Benchmarks of atmospheric nitrogen deposition for several ecosystem indicators 

(REA 5.3.1.2) MCF-Mixed Conifer Forest, CSS-Coastal Sage Scrub.............. 2-25 

Figure 2-9 Observed effects from ambient and experimental atmospheric nitrogen deposition 

loads in relation to using CMAQ 2002 modeling results and NADP monitoring 

data..................................................................................................................... 2-26 

Figure 3-1 Common anthropogenic stressors and the essential ecological attributes they 

affect. ................................................................................................................... 3-3 

Figure 3-2 European maps of eutrophication (left) and acidification (right) which protect 

95% of natural areas in 50x50 km2 European Monitoring and Evaluation 

Programme grid ................................................................................................... 3-9 

Figure 3-3 Representation of the benefits assessment process indicating where some 

ecological benefits may remain unrecognized, unquantified, or unmonetized.. 3-12 

Figure 3-4 Conceptual model showing the relationships among ambient air quality indicators 

and exposure pathways and the resulting impacts on ecosystems, ecological 

responses, effects and benefits to characterize known or anticipated adverse 

effects to public welfare..................................................................................... 3-14 

Figure 3-5 Locations of Eastern U.S. National Parks (Class I areas) relative to deposition of 

Nitrogen and Sulfur in sensitive aquatic areas................................................... 3-15 

Figure 3-6 Location of Western U.S. National Parks (Class I areas) relative to deposition of 

Nitrogen and Sulfur............................................................................................ 3-16 

Figure 3-7 Conceptual model linking ecological indicator (ANC) to affected ecosystem 

services............................................................................................................... 3-25 

vii
Figure 4-1 Routinely operating surface monitoring stations measuring forms of atmospheric 

nitrogen ................................................................................................................ 4-9 

Figure 4-2 Routinely operating surface monitoring stations measuring forms of atmospheric 

sulfur .................................................................................................................. 4-10 

Figure 4-3 Anticipated network of surface based NOY stations based on 2009 network design 

plans. ............................................................................................................... 4-12 

Figure 4-4 Location of approximately 250 National Atmospheric Deposition Monitoring 

(NADP) National Trends Network (NTN) sites illustrating annual ammonium 

deposition for 2005. ......................................................................................... 4-14 

Figure 4-5 2005 CMAQ modeled annual average NOY (ppb). ......................................... 4-21 

Figure 4-6 2005 CMAQ modeled annual average total reduced nitrogen (NHX) (as ug/m3

nitrogen)............................................................................................................. 4-22 

Figure 4-7 2005 CMAQ modeled annual average ammonia, NH3, (as ug/m3

 N) ............... 4-23 

Figure 4-8 2005 CMAQ modeled annual average ammonium, NH4, (as ug/m3

 N)............ 4-24

Figure 4-9 2005 CMAQ modeled annual average SOX, (as ug/m3

 S from SO2 and SO4) .. 4-25 

Figure 4-10 2005 CMAQ modeled annual average SO2 (as ug/m3

 S) .................................. 4-26 

Figure 4-11 2005 CMAQ modeled annual average SO4 (as ug/m3 S). ................................. 4-27 

Figure 4-12 2005 annual average sulfate concentrations based on CASTNET generated by the 

Visibility Information Exchange Web Sysytem (VIEWS)................................ 4-28 

Figure 4-13 2005 annual average sulfate concentrations based on CASTNET generated by the 

Visibility Information Exchange Web Sysytem (VIEWS)................................ 4-29 

Figure 4-14 Annual average 2005 NOy concentrations from reporting stations in AQS ..... 4-30 

Figure 4-15 2005 CMAQ modeled oxidized nitrogen deposition (kgN/Ha-yr).................... 4-31 

Figure 4-16 2005 CMAQ modeled oxidized sulfur deposition (kgS/Ha-yr)......................... 4-32 

Figure 4-17 2005 CMAQ derived annual average ratio of reduced to total nitrogen 

deposition........................................................................................................... 4-33 

Figure 4-18 Three hour average maximum 2005 SO2 concentrations based on the SLAMS 

reporting to EPA’s Air Quality System (AQS) data base.................................. 4-34 

Figure 4-19 Annual average 2005 NO2 concentrations based on the SLAMS reporting to 

EPA’s Air Quality System (AQS) data base...................................................... 4-35 

Figure 4-20 2005 CMAQ derived annual average ratio of (NOY – NO2)/NOY .................... 4-36 

Figure 4-21 Scatter plots of total oxidized nitrogen deposition with average annual NO2, 

HNO3 and NOY concentrations on each 12 km2

 grid based on 2005 CMAQ results 

for the Adirondack region.................................................................................. 4-37 

Figure 4-22 National map highlighting the 9 case study areas evaluated in the REA .......... 4-41 

Figure 5-1 Conceptual design of the form of NOX and SOX secondary standard ................. 5-6 

Figure 5-2 Conceptual design of the form of the standard based on aquatic acidification ... 5-7 

Figure 5-3 Detailed conceptual design of the form of the standard based on aquatic 

acidification........................................................................................................ 5-10 

Figure 5-4 A. The relationship between pH and ANC under equilibrium conditions with 

mineral phase gibbsite. Triangles indicate calculated values while circles indicate 

measurements (Bi and Liu 2001). B. The relationship between precipitation (wet 

and average year), ANC and the risk to exposure of a pH below 5.5 (Gerritsen et 

al. 1996) ............................................................................................................. 5-13 

Figure 5-5 Illustration of a generalized N + S deposition tradeoff curve that is calculated by 

using the FAB approach..................................................................................... 5-19 

viii
Figure 5-6 Relationship between CLs using pre-industrial base cation weathering ([BC]0

* 

=BCo) calculated by MAGIC versus the F-factor methods. .............................. 5-24 

Figure 5-7 Sites of CLs calculated by SSWC that are used in the Policy Assessment 

 analysis............................................................................................................... 5-27 

Figure 5-8 Cumulative distribution of the critical loads for ANC =50 μeq/L considered in the 

analysis (n =5281).............................................................................................. 5-30 

Figure 5-9 Map that illustrates Omernick’s ecoregions (level 3)........................................ 5-33 

Figure 5-10 Map of sensitive ecoregions (red) using binary categorization approach (sensitive 

vs. less-sensitive) with ANC observations between -50 and 200 μeq/L ........... 5-37 

Figure 5-11 Map of sensitive ecoregions based on clusters of log ANC values between -50 and 

200 μeq/L ........................................................................................................... 5-39 

Figure 5-12 Deposition metrics for 90%,75% and 50% of the population for options, 1 (all 

data), 2a (binary classification), and 2b (clusters based on log ANC) ............. 5-44 

Figure 5-13 Comparison of sites that are protected if the nation is not subdivided and if the 

nation is subdivided into a sensitive and non-sensitive category ...................... 5-45 

Figure 5-14 Cumulative distribution of the 90% deposition metric for 

 ecoregions is shown ........................................................................................... 5-46 

Figure 5-15 Examples of N + (SO2 +SO4

-2) deposition tradeoff curves for option 1 (A) and 

 option 2a(B and C)............................................................................................. 4-48 

Figure 5-16 A U.S. map of NH3 deposition (meq/m2

/yr) overlaying ecoregions (level 3) 

boundaries.......................................................................................................... 5-50 

Figure 5-18 Examples of N + (SO2 +SO4

-2) deposition tradeoff curves for option............... 5-51 

Figure 5-19 Examples of NOy + (SO2 +SO4

-2) deposition tradeoff curves for option 1 (A) and 

option 2a(B and C). Deposition metrics that correspond to protection 75% or 90% 

of the ecosystems considering both the high and low value for NHx 

 deposition........................................................................................................... 5-52 

Figure 5-20 2005 T values for each grid cell in the eastern U.S. domain.............................5.58 

Figure 5-21 Schematic Diagram illustrating the procedure for converting deposition tradeoff 

curves of sulfur and nitrogen to atmospheric concentrations of SOx and NOy 5-59 

Figure 5-22 Inter-annual coefficients of variation (CV) of a) NOy and b) SOx T values, 

based on a series of 2002-2005 CMAQ v4.7 simulation................................... 5-60 

Figure 5-23 Relationship between ANC and pH levels......................................................... 5-70 

Figure 5-24 Summary of results from Gerritson et al 1996. Blue back herring larval mortality 

occurs at pH 6.2. The higher the ANC the less likely it is that the pH will dip 

below 6.2............................................................................................................ 5-71 

Figure 5-25 Mean residual number of species per lake for lakes in Ontario, by pH interval. 

The residual number of species for a lake is the deviation of the observed number 

from the number predicted by lake area ........................................................... 5-76 

Figure 5-26 Relationship between ANC and number of fish species present in aquatic 

freshwater ecosystems in Shenandoah National Park........................................ 5-79 

Figure 5-27 Relationship between ANC and number of fish species present in aquatic 

freshwater ecosystems in Shenandoah National Park........................................ 5-80 

Figure 6-1 Benchmarks of atmospheric nitrogen deposition for several 

 ecosystem indicators ............................................................................................ 6-5 

Figure 7-1 2002 CMAQv4.6 annual average SO2 predicted concentrations versus 

observations at CASTNet sites in the eastern domain....................................... 7-17 

ix
Figure 7-2 2002 CMAQv4.6 annual average SO2-

4 predicted concentrations versus 

observations at CASTNet sites in the eastern domain....................................... 7-18 

Figure 7-3 2002 CMAQv4.6 annual average TNO3 predicted concentrations versus 

observations at CASTNet sites in the eastern domain....................................... 7-19 

Figure 7-4 2002 CMAQv4.6 annual average NH4

+

 predicted concentrations versus 

observations at CASTNet sites in the eastern domain....................................... 7-20 

Figure 7-5 2002–2005 Domain-wide average SO4

2- predicted concentrations and 

observations by month at CASTNet Sites in the eastern domain ...................... 7-21 

Figure 7-6 2002–2005 Domain-wide average TNO3 predicted concentrations and 

observations by month at CASTNet sites in the eastern domain....................... 7-22 

Figure 7-7 2002–2005 Domain-wide average NH4

+

 predicted concentrations and 

observations by month at CASTNet sites in the eastern domain....................... 7-23 

Figure 7-8 Comparison of CMAQ predictions and measurements for 12-hour (6am-6pm) 

average NH3 concentrations, with a monitoring cycle of 4 days on and 4days off, 

at a high emission site (Kenansville) and a low emission urban site (Raleigh) in 

North Carolina compared to CMAQ for July 2004 (top) and August 2004 

(bottom), from Dennis et al., 2010..................................................................... 7-24 

Figure 7-9 Unadjusted (left) and PRISM (right) adjusted CMAQ annual wet deposited sulfate 

for 2002.............................................................................................................. 7-25 

Figure 7-10 Unadjusted (left) and PRISM and bias (right) adjusted CMAQ annual wet 

deposition of nitrate (top) and ammonium (bottom).......................................... 7-26 

Figure 7-11 Spatial and interannual variability of inverse deposition transference ratios, 1/TSOx

and 1/TNOy, for Adirondack (top) and Shenandoah case study areas................. 7-27 

Figure 7-12 Summary of inter-annual and emissions sensitivity variability of sulfur and 

nitrogen deposition transference ratios .............................................................. 7-28 

Figure 7-13 Cumulative regional NH3 budget of advection, wet- and dry deposition, 

calculated for an expanding box starting at the high-emitting Sampson County NC 

cell...................................................................................................................... 7-29 

Figure 7-14 Simulated versus observed annual average surface water SO42-, NO3-, ANC, and 

pH during the model calibration period for each of the 44 lakes in the 

Adirondacks Case Study Area ........................................................................... 7-33 

Figure 7-15 Simulated versus observed annual average surface water SO42-, NO3-, ANC, and 

pH during the model calibration period for each of the 60 streams in the 

Shenandoah Case Study Area. The black line is the 1:1 line............................. 7-34 

Figure 8-1 Annual 2002 – 2004 CMAQ derived annual average fraction of ambient 

concentrations (above ) and deposition (below) of individual NOy species 

delineated by the Adirondack and Shenandoah case study areas and the remainder 

of the Eastern U.S. domain ................................................................................ 8-15 

Figure 8-2 Examples of the Relative Abundance of Several NOy Species Measured at Two 

Rural Southeastern Canadian Sites as a Fraction of the Total Measured NOy

Concentration..................................................................................................... 8-16 

Figure 8-3 Annual average fraction of NOy ambient air contributed by NO2 based on 2005 

CMAQ Eastern U.S. simulation at 12 km grid cell resolution .......................... 8-17 

Figure 8-4 Annual average fraction of NOy ambient air contributed by HNO3 based on 2005 

CMAQ Eastern U.S. simulation at 12 km grid cell resolution .......................... 8-18 

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Figure 8-5 Annual average fraction of NOy ambient air contributed by PAN based on 2005 

CMAQ Eastern U.S. simulation at 12 km grid cell resolution .......................... 8-19 

Figure 8-6 Annual 2002 – 2004 CMAQ derived annual average fraction of ambient 

concentrations (above ) and deposition (below) of individual SOx species 

delineated by the Adirondack and Shenandoah case study areas and the remainder 

of the Eastern U.S. domain ............................................................................... 8-20 

Figure 9-1 Spatially interpolated CMAQ estimates of deposition of reduced nitrogen (2002-

2004 average)....................................................................................................... 9-8 

Figure 9-2 Range of NOx-SOx Tradeoff Curves Across Acid Sensitivity Categories for 

Alternative Levels of Reduced Nitrogen Deposition, for a Target ANC of 50 

μeq/L and a Target of Protection for 90 Percent of Water Bodies .................... 9-14 

List of Tables 

Table 1-1 Description of parameters, units and conventions ............................................... 1-5 

Table 2-1 Ecological Effects Associated with Alternative Levels of Acid Neutralizing 

Capacity (ANC) ................................................................................................... 2-8 

Table 2-2 Summary several commonly used acidification models.................................... 2-9 

Table 3-1 Crosswalk between Ecosystem Services and Public Welfare Effects ............... 3-11 

Table 3-2 Count of Impacted Lakes................................................................................... 3-28 

Table 3-3 Present Value and annualized Benefits, Adirondack Region ............................ 3-29 

Table 3-4 Aggregate Annual Benefit Estimates for the Zero-Out Scenario ...................... 3-30 

Table 3-5 Annual participation and value of outdoor (forest related) activity 

 in the northeast................................................................................................... 3-32 

Table 3-6 Summary of Studies of Select Terrestrial Ecosystem Services ......................... 3-33 

Table 3-7 Summary of Values for Current Levels of Services and Changes in 

 Service Levels in $2007..................................................................................... 3-37 

Table 3-8 Summary of Annual Damages to Services due to Atmospheric Loading ......... 3-38 

Table 3-9 Summary of Current Levels of Ecosystem Services.......................................... 3-40 

Table 4-1 Summary of Monitoring Networks...................................................................... 4-7 

Table 5-1 Descriptive statistics of the national CL data set............................................... 5-29 

Table 5-2 Summary of data sources considered for the evaluation of national ANC........ 5-34 

Table 5-3. National ANC data. Observations from 6318 sites remained after filtering data to 

remove sites with an average value of ANC<-50 and >200 μeq/L.................... 5-36 

Table 5-4 Descriptive statistics of the CL populations that result when the US is divided into 

two categories, sensitive and less-sensitive based on ANC data....................... 5-37 

Table 5-5 Descriptive statistics of the CL populations that result when the US is divided into 

5 clusters based on log ANC values between -50 and 200 μeq/L..................... 5-39 

Table 5-6 Descriptive statistics of the populations of CL ANC=50 μeq/L values that result 

for each ecoregion.............................................................................................. 5-40 

Table 5-7 Summary of the number of deposition metrics that would result from the options 

for how to categorized the landscape based on acid-sensitivity ....................... 5-43 

Table 5-9 Summary of Fish Mortality Response to pH ..................................................... 5-73 

Table 5-10 Threshold response of increased mortality of fish to low pH listed from least 

 sensitive to most sensitive.................................................................................. 5-74 

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Table 5-11 General summary of biological changes anticipated with surface water 

acidification, expressed as a decrease in surface water pH ............................... 5-77 

Table 5-12 Comparison of percentage protection from ANC values less than 50 and less than 

20 using DL that result when the US is considered one population .................. 5-88 

Table 5-13 Comparison of percentage protection from ANC values less than 50 and less than 

20 using DL that result when the US is divided into two categories, sensitive and 

less sensitive based on ANC data ...................................................................... 5-89 

Table 6-1 Results of the comparison of lake and stream aquatic critical loads (ANC of 50 

μeq/L) to terrestrial critical loads (Bc:Al molar ratios of 10.0 in soil solution) 

calculated for the full watershed in each of the 16 watersheds in the Adirondacks 

and Shenandoah Case Study Areas...................................................................... 6-3 

Table 6-2 Results of the comparison of lake and stream aquatic critical loads (ANC of 50 

μeq/L) to terrestrial critical loads (Bc:Al molar ratios of 1.2 in soil solution) 

calculated for the full watershed in each of the 16 watersheds in the Adirondacks 

and Shenandoah Case Study Areas...................................................................... 6-3 

Table 7-1 Normalized Mean Bias Statistics for Predicted and Observed Pollutant 

Concentration..................................................................................................... 7-11 

Table 7-2 Summary (Incomplete) of Qualitative Uncertainty Analysis of Key Modeling 

Elements in the NOx/SOx AAPI ....................................................................... 7-38 

xii
 List of Acronyms and Abbreviations

AAPI Atmospheric Acidification Potential Index 

ADR Adirondack Mountains of New York 

Al3+ aluminum 

ANC acid neutralizing capacity 

AQCD Air Quality Criteria Document 

AQRV air quality related values 

ASSETS EI Assessment of Estuarine Trophic Status eutrophication index 

Bc/Al Base cation to aluminum ratio, also Bc:Al 

C carbon 

Ca/Al calcium to aluminum ratio 

Ca2+ calcium 

CAA Clean Air Act 

CASAC Clean Air Scientific Advisory Committee 

CASTNet Clean Air Status and Trends Network 

CCS coastal sage scrub 

Chl a chlorophyll a 

CLE critical load exceedance 

CMAQ Community Multiscale Air Quality model 

CSS coastal sage scrub 

CWA Clean Water Act 

DIN dissolved inorganic nitrogen 

DO dissolved oxygen 

DOI U.S. Department of Interior 

EMAP Environmental Monitoring and Assessment Program 

EPA U.S. Environmental Protection Agency 

FHWAR fishing, hunting and wildlife associated recreation survey 

FIA Forest Inventory and Analysis National Program 

FWS Fish and Wildlife Service 

GIS geographic information systems 

GPP gross primary productivity 

H+ hydrogen ion 

H2O water vapor 

H2SO4 sulfuric acid 

ha hectare 

HAB harmful algal bloom 

HFC hydrofluorocarbon 

Hg+2 reactive mercury 

Hg0 elemental mercury 

HNO3 nitric acid

HONO nitrous acid 

HUC hydrologic unit code 

IMPROVE Interagency Monitoring of Protected Visual Environments 

ISA Integrated Science Assessment 

K+

 potassium 

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kg/ha/yr kilograms per hectare per year 

km kilometer 

LRMP Land and Resource Management Plan 

LTER Long Term Ecological Monitoring and Research 

LTM Long-Term Monitoring 

MAGIC Model of Acidification of Groundwater in Catchments 

MCF Mixed Conifer Forest 

MEA Millennium Ecosystem Assessment 

Mg2+ magnesium 

N nitrogen 

N2 gaseous nitrogen 

N2O nitrous oxide 

N2O3 nitrogen trioxide 

N2O4 nitrogen tetroxide 

N2O5 dinitrogen pentoxide 

Na+ sodium 

NAAQS National Ambient Air Quality Standards 

NADP National Atmospheric Deposition Program 

NAPAP National Acid Precipitation Assessment Program 

NAWQA National Water Quality Assessment 

NEEA National Estuarine Eutrophication Assessment 

NEP net ecosystem productivity 

NH3 ammonia gas 

NH4

+

 ammonium ion 

(NH4)2SO4 ammonium sulfate 

NHx category label for NH3 plus NH4

+ 

NO nitric oxide 

NO2 nitrogen dioxide 

NO2

- reduced nitrite 

NO3

- reduced nitrate 

NOAA National Oceanic and Atmospheric Administration 

NOx nitrogen oxides 

NOy total oxidized nitrogen 

NPP net primary productivity 

NPS National Park Service 

NRC National Research Council 

NSWS National Surface Water Survey 

NTN National Trends Network 

NTR organic nitrate 

O3 ozone 

OAQPS Office of Air Quality Planning and Standards 

OW Office of Water 

PAN peroxyacyl nitrates 

PFC perfluorocarbons 

pH relative acidity 

ppb parts per billion 

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ppm parts per million 

ppt parts per trillion 

PSD prevention of significant deterioration 

REA Risk and Exposure Assessment 

REMAP Regional Environmental Monitoring and Assessment Program 

S sulfur 

S2O3 thiosulfate 

S2O7 heptoxide 

SAV submerged aquatic vegetation 

SF6 sulfur hexafluoride 

SMP Simple Mass Balance 

SO sulfur monoxide 

SO2 sulfur dioxide 

SO3 sulfur trioxide 

SO3

2- sulfite 

SO4 wet sulfate 

SO4

2- sulfate ion 

SOM soil organic matter 

SOx sulfur oxides 

SPARROW SPAtially Referenced Regressions on Watershed Attributes 

SRB sulfate-reducing bacteria 

STORET STORage and RETrieval 

TIME Temporally Integrated Monitoring of Ecosystems 

TMDL total maximum daily load 

TP total phosphorus 

USFS U.S. Forest Service 

USGS U.S. Geological Survey 

μeq/L microequivalents per liter 

g/m3 micrograms per cubic meter 

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List of Key Terms 

Acidification: The process of increasing the acidity of a system (e.g., lake, stream, forest soil). 

Atmospheric deposition of acidic or acidifying compounds can acidify lakes, streams, 

and forest soils. 

Air Quality Indicator: The substance or set of substances (e.g., PM2.5, NO2, SO2) occurring in 

the ambient air for which the National Ambient Air Quality Standards set a standard level 

and monitoring occurs. 

Alpine: The biogeographic zone made up of slopes above the tree line, characterized by the 

presence of rosette-forming herbaceous plants and low, shrubby, slow-growing woody 

plants. 

Acid Neutralizing Capacity: A key indicator of the ability of water to neutralize the acid or 

acidifying inputs it receives. This ability depends largely on associated biogeophysical 

characteristics, such as underlying geology, base cation concentrations, and weathering 

rates. 

Arid Region: A land region of low rainfall, where “low” is widely accepted to be less than 250 

mm precipitation per year. 

Base Cation Saturation: The degree to which soil cation exchange sites are occupied with base 

cations (e.g., Ca2+, Mg2+, K+) as opposed to Al3+ and H+. Base cation saturation is a 

measure of soil acidification, with lower values being more acidic. There is a threshold 

whereby soils with base saturations less than 20% (especially between 10%–20%) are 

extremely sensitive to change. 

Ecologically Relevant Indicator: A physical, chemical, or biological entity/feature that 

demonstrates a consistent degree of response to a given level of stressor exposure and 

that is easily measured/quantified to make it a useful predictor of ecological risk. 

Critical Load: A quantitative estimate of an exposure to one or more pollutants, below which 

significant (as defined by the analyst or decision maker) harmful effects on specified 

sensitive elements of the environment do not occur, according to present knowledge. 

Denitrification: The anaerobic reduction of oxidized nitrogen (e.g., nitrate or nitrite) to gaseous 

nitrogen (e.g., N2O or N2) by denitrifying bacteria. 

Dry Deposition: The removal of gases and particles from the atmosphere to surfaces in the 

absence of precipitation (e.g., rain, snow) or occult deposition (e.g., fog). 

Ecological Risk: The likelihood that adverse ecological effects may occur or are occurring as a 

result of exposure to one or more stressors (U.S. EPA, 1992). 

Ecological Risk Assessment: A process that evaluates the likelihood that adverse ecological 

effects may occur or are occurring as a result of exposure to one or more stressors (U.S. 

EPA, 1992). 

Ecosystem: The interactive system formed from all living organisms and their abiotic (i.e., 

physical and chemical) environment within a given area. Ecosystems cover a hierarchy of 

spatial scales and can comprise the entire globe, biomes at the continental scale, or small, 

well-circumscribed systems such as a small pond. 

Ecosystem Benefit: The value, expressed qualitatively, quantitatively, and/or in economic terms, 

where possible, associated with changes in ecosystem services that result either directly 

or indirectly in improved human health and/or welfare. Examples of ecosystem benefits 

that derive from improved air quality include improvements in habitats for sport fish 

species, the quality of drinking water and recreational areas, and visibility. 

xvi
Ecosystem Function: The processes and interactions that operate within an ecosystem. 

Ecosystem Services: The ecological processes or functions having monetary or non-monetary 

value to individuals or society at large. These are (1) supporting services, such as 

productivity or biodiversity maintenance; (2) provisioning services, such as food, fiber, or 

fish; (3) regulating services, such as climate regulation or carbon sequestration; and (4) 

cultural services, such as tourism or spiritual and aesthetic appreciation. 

Eutrophication: The process by which nitrogen additions stimulate the growth of autotrophic 

biota, usually resulting in the depletion of dissolved oxygen. 

Nitrogen Enrichment: The process by which a terrestrial system becomes enhanced by nutrient 

additions to a degree that stimulates the growth of plant or other terrestrial biota, usually 

resulting in an increase in productivity. 

Nitrogen Saturation: The point at which nitrogen inputs from atmospheric deposition and other 

sources exceed the biological requirements of the ecosystem; a level beyond nitrogen 

enrichment. 

Occult Deposition: The removal of gases and particles from the atmosphere to surfaces by fog 

or mist. 

Semi-arid Regions: Regions of moderately low rainfall, which are not highly productive and are 

usually classified as rangelands. “Moderately low” is widely accepted as between 100- 

and 250-mm precipitation per year. 

Sensitivity: The degree to which a system is affected, either adversely or beneficially, by NOx 

and/or SOx pollution (e.g., acidification, nutrient enrichment). The effect may be direct 

(e.g., a change in growth in response to a change in the mean, range, or variability of 

nitrogen deposition) or indirect (e.g., changes in growth due to the direct effect of 

nitrogen consequently altering competitive dynamics between species and decreased 

biodiversity). 

Total Reactive Nitrogen: This includes all biologically, chemically, and radiatively active 

nitrogen compounds in the atmosphere and biosphere, such as NH3, NH4+, NO, NO2, 

HNO3, N2O, NO3–, and organic compounds (e.g., urea, amines, nucleic acids). 

Valuation: The economic or non-economic process of determining either the value of 

maintaining a given ecosystem type, state, or condition, or the value of a change in an 

ecosystem, its components, or the services it provides. 

Variable Factors: Influences which by themselves or in combination with other factors may 

alter the effects on public welfare of an air pollutant (section 108 (a)(2)) 

 (a) Atmospheric Factors: Atmospheric conditions that may influence transformation, 

conversion, transport, and deposition, and thereby, the effects of an air pollutant on 

public welfare, such as precipitation, relative humidity, oxidation state, and co-pollutants 

present in the atmosphere. 

 (b) Ecological Factors: Ecological conditions that may influence the effects of an air 

pollutant on public welfare once it is introduced into an ecosystem, such as soil base 

saturation, soil thickness, runoff rate, land use conditions, bedrock geology, and 

weathering rates. 

Vulnerability: The degree to which a system is susceptible to, and unable to cope with, the 

adverse effects of NOx and/or SOx air pollution. 

Welfare Effects: The effects on soils, water, crops, vegetation, man-made materials, animals, 

wildlife, weather, visibility, and climate; as well as damage to and deterioration of 

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property, hazards to transportation, and the effects on economic values and on personal 

comfort and well-being, whether caused by transformation, conversion, or combination 

with other air pollutants (Clean Air Act Section 302[h]). 

Wet Deposition: The removal of gases and particles from the atmosphere to surfaces by rain or 

other precipitation. 
1 1 INTRODUCTION 

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The U.S. Environmental Protection Agency (EPA) is presently conducting a review of 

the secondary National Ambient Air Quality Standards (NAAQS) for oxides of nitrogen (NOx) 

and oxides of sulfur (SOx). The EPA’s overall plan and schedule for this review were presented 

in the Integrated Review Plan for the Secondary National Ambient Air Quality Standards for 

Nitrogen Dioxide and Sulfur Dioxide (US EPA, 2007). The Integrated Review Plan (IRP) 

outlined the Clean Air Act (CAA or the Act) requirements related to the establishment and 

reviews of the NAAQS, the process and schedule for conducting the current review, and the key 

components in the NAAQS review process: an Integrated Science Assessment (ISA), Risk and 

Exposure Assessment (REA), and policy assessment/rulemaking. It presented key policyrelevant issues to be addressed in this review as a series of questions that frames our 

consideration of whether the current secondary (welfare-based) NAAQS for NOx and SOx 

should be retained or revised. 

As part of this review, staff in the U.S. Environmental Protection Agency’s (EPA) Office 

of Air Quality Planning and Standards (OAQPS) prepared this second draft Policy Assessment.1

 

The objective of this assessment is to evaluate the policy implications of the key scientific 

information contained in the document Integrated Science Assessment for Oxides of Nitrogen 

and Sulfur-Ecological Criteria (USEPA, 2008; henceforth referred to as the ISA), prepared by 

EPA’s National Center for Environmental Assessment (NCEA) and the results from the analyses 

contained in the Risk and Exposure Assessment for Review of the Secondary National Ambient 

Air Quality Standards for Oxides of Nitrogen and Oxides of Sulfur (U.S. EPA, 2009; henceforth 

referred to as the REA). This second draft also presents staff conclusions on a range of policy 

options that we believe are appropriate for the Administrator to consider concerning whether, 

and if so how, to revise the secondary (welfare-based) NOx and SOx NAAQS. 

1

 Preparation of a PA by OAQPS staff reflects Administrator Jackson’s decision to modify the NAAQS review 

process that was presented in the IRP. See http://www.epa.gov/ttn/naaqs/review.html for more information on the 

current NAAQS review process. 

September, 2010 1-1 Draft – Do Not Quote or Cite .
This policy assessment is intended to help “bridge the gap” between the scientific 

assessment contained in the ISA and the judgments required of the EPA Administrator in 

determining whether it is appropriate to retain or revise the secondary NAAQS for NOx and 

SOx. This policy assessment considers the available scientific evidence and quantitative riskbased analyses, together with related limitations and uncertainties, and focuses on the basic 

components of air quality standards: indicators

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, averaging times, forms3, and levels. These 

components, which serve to define each standard, must be considered collectively in evaluating 

the welfare protection afforded by the secondary NOx and SOx NAAQS. Our development of 

this policy assessment is based on the assessment and integrative synthesis of information 

presented in the ISA and on staff analyses and evaluations presented in this document, and is 

further informed by comments and advice received from an independent scientific review 

committee, the Clean Air Scientific Advisory Committee (CASAC), in their review of the 

previous integrated science assessment, risk and exposure assessment, and first draft policy 

assesment. To view related documents developed as part of the planning, science, and risk 

assessment phases of this review see 

http://www.epa.gov/ttn/naaqs/standards/no2so2sec/index.html. 

This document is organized around a conceptual framework for a combined NOx and SOx 

secondary NAAQS and is focused on answering key policy questions related to the 

implementation of that conceptual framework. Chapter 2 provides a summary of ecological 

effects from the deposition of ambient NOx and SOx to sensitive ecosystems, drawing from the 

ISA and REA. Chapter 3 places those ecological effects within the context of “public welfare” 

by linking effects to ecosystem services or other benchmarks of public welfare. Chapter 4 

addresses the adequacy of the current NOx and SOx secondary NAAQS in addressing the 

impacts on public welfare from ecological effects. Chapter 5 develops the conceptual design for 

ecologically relevant multi-pollutant standards and presents options for developing critical 

components of a secondary NAAQS necessary to implement the conceptual design. Chapter 6 

describes how secondary NAAQS designed to protect a specific ecological endpoint may also 

provide protection for other ecological endpoints. Chapter 7 provides an assessment of critical 

2

 The “indicator” of a standard defines the chemical species or mixture that is to be measured in determining 

whether an area attains the standard. 

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 The “form” of a standard defines the air quality statistic that is to be compared to the level of the standard in 

determining whether an area attains the standard.

September 2010 1-2 Draft –Do Not Quote or Cite 
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uncertainties and sensitivitie considered in developing the options for the components of the 

standard. Chapter 8 discusses issues related to monitoring of NOx and SOx. Chapter 9 provides 

initial staff conclusions regarding options for pollutant indicators, averaging times, forms, and 

ranges of levels for the secondary NOx and SOx NAAQS. 

 In this document we consider how the available scientific evidence and quantitative riskbased analyses, together with related limitations and uncertainties, inform the review of each 

element of the NAAQS: indicator, averaging times, forms, and levels. These components must be 

considered collectively in evaluating the welfare protection afforded by the secondary NAAQS

standards. This draft document does not contain final staff conclusions as to all the necessary 

components of an alternative secondary standard for NOx and/or SOx but rather describes the 

current state of thinking with regard to potential policy options and provides an appropriate 

context of information for the Administrator to consider in making decisions regarding the 

standards. 

While this policy assessment should be of use to all parties interested in the secondary 

NOx and SOx NAAQS review, it is written with an expectation that the reader has some 

familiarity with the technical discussions contained in the ISA and REA. 

EPA’s final Policy Assessment will address additional CASAC comments on this second 

draft, and will include sufficient information to inform the Administrator on critical components 

of the standards, and staff conclusions regarding alternative levels of the standards. 

21 1.1 DEFINITIONS, PARAMETERS, UNITS, AND CONVENTIONS USED FOR THIS 

ASSESSMENT 

Throughout this document numerous terms are used that address a variety of atmospheric 

and ecosystem processes and variables. Some of the more common terms used in the technical 

community are not always synonymous with definitions imbedded in the CAA. Because of this 

diversity of terms spanning atmospheric and ecosystem processes along with adherence to 

scientific and legal conventions, this section provides the terminology as a reference source for 

the entire report. 

As discussed in detail in the REA (REA 1.3.1), in the atmospheric science community 

NOx is typically referred to as the sum of nitrogen dioxide (NO2), and nitric oxide (NO). As 

defined by the Clean Air Act, the family of NOx includes any gaseous combination of nitrogen 

September 2010 1-3 Draft –Do Not Quote or Cite 
and oxygen (e.g., NO2, NO, nitrous oxide [N2O], dinitrogen trioxide [N2O3], dinitrogen tetroxide 

[N2O4], and dinitrogen pentoxide [N2O5]). The term used by the scientific community to 

represent the complete set of reactive oxidized nitrogen compounds, including those listed in 

CAA Section 108(c) with the exception of N2O, is total oxidized nitrogen (NOy), commonly 

defined as NO, NO2 and the all of the oxidation products of NO and NO2 Reactive oxidized 

nitrogen is defined as NOy = NO2 + NO + HNO3 + PAN +2N2O5 + HONO+ NO3 + organic 

nitrates + particulate NO3 (Finlayson-Pitts and Pitts, 2000). In this document, unless otherwise 

indicated, we use the term NOy as the atmospheric indicators associated with the NOx 

component of the proposed NOx/SOx standard . 

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For this assessment, SOx is defined to include all oxides of sulfur, including multiple 

gaseous substances (e.g., SO2, sulfur monoxide [SO], sulfur trioxide [SO3], thiosulfate [S2O3], 

and heptoxide [S2O7], as well as particulate species, such as ammonium sulfate [(NH4)2SO4]). 

Throughout this text we refer to sulfate as SO4 and nitrate as NO3, recognizing that they have 

charges of -2 for sulfate and -1 for nitrate. The sum of sulfur dioxide gas (SO2) and particulate 

sulfate (SO4), referred herein as (SO2 + SO4) is used throughout this document as the 

atmospheric indictor for the SOx component of the proposed NOx/SOx standard. From a 

measurement and modeling perspective we only consider the sum of SO2 and particulate SO4 as 

the indicator for sulfur. The sum of SO2 and SO4 constitute virtually all of the ambient air sulfur 

budget and are measured routinely in monitoring networks. 

 Table 1-1 provides further explanation of these indicators, some of which is repeated in 

Chapters 4, 5 and 8. Table 1 also provides details on the units used throughout the equations 

and examples in the PAD. Again, because of difference in unit conventions between 

atmospheric and ecosystem sciences, there are detailed explanations of units as well as 

procedures for translating between different unit conventions. To facilitate the linkage between 

atmospheric and ecosystem processes, only the mass (or equivalent charge) associated with 

sulfur or nitrogen is considered in mass, mixing ratio, and deposition unit conventions. 

 

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1 Table 1 -1. Description of parameters, units and conventions.

Parameter Units Conversions to other 

unit conventions used in 

figures and calculations 

(multiply value in Units 

column by: 

Explanation 

Atmospheric species 

CMAQ defined NOy species: NO (nitrogen oxide), NO2 (nitrogen dioxide), HNO3 (nitric acid), p-NO3 (particulate 

bound nitrate), NO3 (sum of HNO3 and p-NO3), PAN (peroxy acetyl nitrate), N2O5 (dinitrogen pentoxide), PANX 

(higher order PANs), NTR (organic nitrates), PNA (HNO4); sulfur dioxide (SO2), particulate sulfate (SO4); 

NHx species: NH3 (ammonia), ammonium ion (NH4)

Lumped Atmospheric Species 

NOy The sum of all reactive oxidized nitrogen compounds derived through 

summing all nitrogen contributions (i.e., 1·HNO3 + 2·N205 +...) from the 

modeled species (HNO3, p-NO3, NO2, NO, PAN, ...) or through direct 

measurement which reduces all oxidized nitrogen species to NO and reports 

as ppb NO. All references to the quantity NOy refer to the mass, molar or 

equivalent charge contribution of nitrogen only. All mass contributions of 

oxygen, hydrogen and carbon are not included.

(SO2 +SO4) Oxidized forms of sulfur defined as sulfate (SO4 + SO2); mass units 

maintained for consistency with deposition calculations 

Note that only mass as sulfur is counted in state variables; in practice, 

individual SO2 and SO4 are measured/modeled and converted to mass of 

sulfur atoms or equivalent charge units. Mass contribution of oxygen is 

not included. 

NHx Reduced nitrogen calculated as the sum of NH3 and NH4. All references to 

the quantity NHx used as state variables refer to the mass, molar or 

equivalent charge contribution of nitrogen only. Mass contribution of 

oxygen is not included.

Atmospheric State Variables used in equations and derivations 

NOy concentration 

SOx

NHx 

Used in various 

conventions of: 

Ci; 

μg/m3

 as N or S ppb = (MA/Mi)· ρair) ·μg/m3

where ρair is the air density 

in units of (kg/m3

); 

ρair = 28.97( 10)-3·P/(R·T) 

R = 8.206(10)-

5

m3atm/(mol·K) 

P = atm 

T = degrees K 

MA = molecular weight of 

air (28.97) 

Mi = Atomic weight of 

nitrogen (14) or sulfur (32) 

meq/m3

 = (1/Mi) ·μg/m3 

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Parameter Units Conversions to other 

unit conventions used in 

figures and calculations 

(multiply value in Units 

column by: 

Explanation 

NOy deposition 

(SO2 +SO4)

NHx 

Used in various 

conventions of: 

Depi

meq/m2

-yr as N 

or S

kg/ha-yr =(Mi/q)(10)-2 ·

meq/m2

-yr 

where q = charge (1 for N, 2 

for S) 

max C SOSO )24( μg/m3 is the concentration of 

(SO2 +SO4) in the 

atmosphere consistent 

with max DLS

max CNOy μg/m3 is the concentration of 

NOy in the atmosphere 

consistent with max DLN

min CNOy μg/m3 is the concentration of 

NOy in the atmosphere 

consistent with min DLN

v

Dry

i

m/yr dry deposition velocities 

v

Wet

i

m/yr wet deposition velocities 

DepDry

i

meq/m2

-yr dry deposition fluxes 

DepWet

i

meq/m2

-yr wet deposition fluxes 

Total Depi meq/m2

-yr total (wet+dry) 

deposition 

TSOx 

TNOy

m/yr Calculated by dividing total 

((SO2 +SO4) or NOy 

deposition (wet and dry) by 

the annual average (SO2

+SO4) or NOy 

concentration. 

the transfer ratio, which 

can be considered an 

aggregated, “effective” 

deposition velocity that 

relates total deposition 

of (SO2 +SO4) or NOy to 

the total ambient 

concentration, and 

represents an average of 

the chemical species 

specific vi

Tot ( = viDry + 

vi

Wet) values 

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Parameter Units Conversions to other 

unit conventions used in 

figures and calculations 

(multiply value in Units 

column by: 

Explanation 

Ecosystem variables 

ANC* μeq/L calculated value of ANC 

ANClimit μeq/L a “target” ANC level 

(μeq/L) 

CLANClim(i) meq/m2

-yr depositional load that 

does not cause the 

catchment to exceed a 

given ANClim, where i

indicates the pollutant of 

interest 

Q m/yr Average surface water 

runoff for an acid 

sensitive area (this is 

typically equal to 

precipitation –

evapotranspiration 

Qws m/yr Catchment level surface 

water runoff (m/yr) (this 

is typically equal to 

precipitation –

evapotranspiration 

NECO meq/m2

-yr nitrogen retention and 

denitrification by 

terrestrial catchment 

Nleach meq/m2

-yr N leaching 

DL%eco(i) meq/m2

-yr The deposition metric, 

defined as the amount of 

deposition that protects 

a selected percentage of 

individual catchments 

for a population of 

water bodies from 

exceeding their 

DLANClim(i), where i

indicates the pollutant 

of interest max DLN meq/m2

-yr In the tradeoff curve 

for DL%eco(i), the 

maximum of amount 

of N deposition when 

S deposition equals 

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Parameter Units Conversions to other 

unit conventions used in 

figures and calculations 

(multiply value in Units 

column by: 

Explanation 

zero 

min DLN

meq/m2

-yr In the tradeoff curve 

for DL%eco(i), the 

amount of N 

deposition that will be 

captured by the 

ecosystem before it 

leaches 

max DLS

meq/m2

-yr In the tradeoff curve 

for DL%eco(i), the 

maximum amount of S 

sulfur deposition 

considering NECO 

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1.2 POLICY OBJECTIVES 

In conducting this periodic review of the NOx and SOx secondary NAAQS, EPA has 

decided to jointly assess the scientific information, associated risks, and standards relevant to 

protecting the public welfare from adverse effects associated with oxides of nitrogen and sulfur. 

Although EPA has historically adopted separate secondary standards for oxides of nitrogen 

(NOx) and oxides of sulfur (SOx), EPA is conducting a joint secondary review of these standards 

because NOx, SOx, 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). There is a strong basis 

for considering these pollutants together, building upon EPA’s and CASAC’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. 

EPA sets secondary standards for two criteria pollutants related to NOx and SOx: ozone 

and particulate matter (PM). NOx is a precursor 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. SOx is a precursor to the formation of particulate sulfate, 

which is a significant component of fine PM in many parts of the U.S. While there are a number 

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of welfare effects associated with ozone and fine PM, including ozone damage to vegetation, and 

visibility degradation related to PM, protection against those effects is provided by the ozone and 

fine PM standards. This review focuses on evaluation of the protection provided by NOx and 

SOx secondary standards for effects associated with direct atmospheric concentrations of NOx 

and SOx, and effects associated with deposition of NOx and SOx to sensitive ecosystems, 

including deposition in the form of particulate nitrate and sulfate in their component forms. 

The ISA highlights the ecological effects associated with deposition of ambient NOx and 

SOx to sensitive ecosystems other than commercially managed forests and agricultural lands. 

This assessment primarily focuses on the effects of ambient NOx and SOx via deposition on 

multiple ecological receptors, but also evaluates information on gas-phase effects of NOx and 

SOx via stomatal exposure on vegetation, which are the effects that the current secondary 

standards protect against. The ISA highlighted effects including those associated with 

acidification and nitrogen nutrient enrichment. Based on these highlighted effects, EPA’s policy 

objective is to develop a framework for NOx and SOx standards that incorporate factors that will 

lead to standards that are ecologically relevant, and that recognizes the interactions between the 

two pollutants as they deposit to sensitive ecosystems, with an ultimate goal of setting standards 

that, based on the ecological criteria described in the ISA, and consistent with the requirements 

of the Clean Air Act, “are requisite to protect the public welfare from any known or anticipated 

adverse effects associated with the presence of such air pollutant in the ambient air.” 

 In presenting policy options for the Administrator’s consideration, we note that the final 

decision on retaining or revising the current secondary standards for NOx and SOx is largely a 

public welfare policy judgment based on the Administrator’s informed assessment of what 

constitutes requisite protection against adverse effects to public welfare. A final 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 in 

ecosystem services. Our approach to informing these judgments, as discussed below, is 

consistent with the requirements of the NAAQS provisions of the Clean Air Act and with how 

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EPA and the courts have historically interpreted the Act. These provisions require the 

Administrator to establish secondary NAAQS that, in the Administrator’s judgment, are 

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

the presence of NOx and SOx in the ambient air. In so doing, the Administrator seeks to 

establish standards that are neither more nor less stringent than necessary for this purpose. 

 For this second draft policy assessment, we have chosen to focus much of our discussion 

on the effects of ambient NOx and SOx on ecological impacts associated with acidifying 

deposition of nitrogen and sulfur, which is a transformation product of ambient NOx and SOx. 

We have the greatest confidence in the causal linkages between NOx and SOx and aquatic 

acidification effects, and we have the most complete information available with which to develop 

an ecologically meaningful structure for the standards. 

1.3 CRITICAL POLICY ELEMENTS 

Our policy objective is guided by the information in the ISA and REA, framed within the 

legislative requirements of the CAA. This framing leads us to focus on critical policy elements 

(CPE) consistent with elements of Clean Air Act language. 

Sections 108 and 109 of the CAA govern the establishment and periodic review of the 

NAAQS and of the air quality criteria upon which the standards are based. The NAAQS are 

established for pollutants that are listed under section 108, based on three criteria, including 

whether emissions of the air pollutant cause or contribute to air pollution which may reasonably 

be anticipated to endanger public health or welfare and whose presence in the ambient air results 

from numerous or diverse mobile or stationary sources. The NAAQS are based on air quality 

criteria that reflect the latest scientific knowledge, useful in indicating the types and extent of 

identifiable effects on public health or welfare that may be expected from the presence of the 

pollutant in ambient air. The criteria refer to criteria issued pursuant to §108 of the Clean Air 

Act, which include “(A) those variable factors (including atmospheric conditions) which of 

themselves or in combination with other factors may alter the effects on public health or welfare 

of such air pollutant; (B) the types of air pollutants which, when present in the atmosphere, may 

interact with such pollutant to produce an adverse effect on public health of welfare; and (C) any 

known or anticipated adverse effects on welfare.” 

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The following critical policy elements for the design of ecologically relevant secondary 

standards for NOx and SOx are identified: 

 (CPE 1) An evaluation of the effects of ambient NOx and SOx on ecosystems, and the 

relationship between those effects and the measure of dose in the ecosystem, 

indicated by the depositional loadings of N and S. 

 (CPE 1.1) Evaluation of the relationship between response of ecological 

receptors, e.g. changes in diversity of fish species, and the response related to 

public welfare, e.g. loss in recreational fishing services. 

 (CPE 1.2) Evaluation of the extent to which identified effects are occurring 

under recent conditions, and the extent to which meeting the current standards 

would provide protection against these effects. 

 (CPE 2) An assessment of how best to characterize, in defining the standards, the 

variable ecosystem factors that affect the relationship between ecological 

effects and depositional loadings of N and S. 

 (CPE 2.1) Specification of potential indicators of ecological effects, e.g. acid 

neutralizing capacity (ANC) that incorporates variability in ecosystem factors. 

 (CPE 3) Characterization of the complex relationships between ambient concentrations 

of NOx and SOx and deposition of N and S in the specification of a standard. 

 (CPE 4) Specification of the form for the standard(s), including ambient atmospheric 

indicators for NOx and SOx, with consideration of averaging times, and 

options for levels of the standard(s). 

The development of the conceptual framework for the NOx and SOx standards described 

in Section 1.4 will be motivated by these critical policy elements. However, in order to provide a 

historical context for this new framework, the next section provides a brief history of previous 

reviews of the NOx and SOx secondary NAAQS, as well as other relevant historical reviews of 

welfare effects associated with these pollutants. 

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1.4 HISTORICAL CONTEXT 

1.4.1 History of NOx and SOx NAAQS Review 

1.4.1.1 NOx NAAQS 

EPA began the most recent previous review of the NOx secondary standards in 1987 and in 

November 1991, EPA released an updated draft AQCD for CASAC and public review and 

comment (56 FR 59285). This draft document provided a comprehensive assessment of the 

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

and other NOx. 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 Air Quality Criteria for Oxides of Nitrogen was then finalized (U.S. 

EPA, 1993). EPA also prepared a Staff Paper that summarized and integrated the key studies 

and scientific evidence contained in the revised NOx AQCD and identified the critical elements 

to be considered in the review of the NO2 NAAQS. 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). The level 

for both the existing primary and secondary NAAQS for NO2 is 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. 

1.4.1.2 SOx NAAQS 

Based on the 1970 SOx criteria document (DHEW, 1970), EPA promulgated primary and 

secondary NAAQS for SO2 on April 30, 1971 (36 FR 8186). The secondary standards included 

a standard at 0.02 ppm in an annual arithmetic mean and a 3-hour average of 0.5 ppm, not to be 

exceeded more than once per year. These secondary standards were established solely on the 

basis of evidence of adverse effects on vegetation. In 1973, revisions made to Chapter 5 

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(“Effects of Sulfur Oxide in the Atmosphere on Vegetation”) of Air Quality Criteria for Sulfur 

Oxides (U.S. 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 SOx air concentrations have 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 SOx concentrations and effects (38 FR 

25679). 

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In 1979, EPA announced that it was revising the Air Quality Criteria Document (AQCD) 

for sulfur oxides concurrently with that for particulate matter and would produce a combined 

particulate matter and sulfur oxides 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. 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. 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. CASAC also suggested that a discussion of acid 

deposition be included in the AQCDs for nitrogen oxides and PM and SOx. Following CASAC 

closure on the AQCD for SO2 in December 1981, EPA’s Office of Air Quality Planning and 

Standards published a Staff Paper in November 1982, but the paper did not directly assess the 

issue of acid deposition. Instead, EPA subsequently prepared the following documents: The 

Acidic Deposition Phenomenon and Its Effects: Critical Assessment Review Papers, Volumes I 

and II (U.S. EPA, 1984a, b), and The Acidic Deposition Phenomenon and Its Effects: Critical 

Assessment Document (U.S. EPA, 1985) (53 FR 14935 -14936). These documents, though they 

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were not considered criteria documents and did not undergo CASAC review, represented the 

most comprehensive summary of relevant scientific information completed by EPA at that point. 

On April 26, 1988 (53 FR 14926), EPA proposed not to revise the existing primary and 

secondary standards for SO2 . This proposal regarding 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. 

1.4.2 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 NOx emissions 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. 

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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 (U.S. EPA, 1995) in 

fulfillment of this requirement. The 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.” 

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 increasingdamage 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.” 

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

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

warming, and tropospheric ozone and stratospheric ozone depletion. 

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In a related matter, the Office of the Secretary of the U.S. Department of Interior 

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, EPA announced receipt of these requests and asked 

for comment on the issues raised in them. 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 (65 FR 

48699). 

The most recent 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).4

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

the NAPAP’s above, EPA believes it is appropriate, 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 and the feasibility of setting a secondary NAAQS to address remaining known 

or anticipated adverse public welfare effects resulting from the acidic and nutrient deposition of 

these criteria pollutants. 

4

 Note that a new NAPAP report is expected to be released later in 2010. The findings of that report will be 

considered in the final policy assessment. 

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1 1.5 PROPOSED CONCEPTUAL FRAMEWORK FOR COMBINED NOX SOX 

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STANDARDS 

There is a strong basis for considering NOx and SOx together at this time, building upon 

EPA’s and CASAC’s 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. The REA introduced a conceptual framework for ecologically meaningful 

secondary standards that recognized the complex processes by which ecosystems are exposed to 

ambient NOx and SOx. That framework provided a flow from ambient concentrations exposures 

via deposition to ecological indicators and effects (see Figure ES-2 in the REA Executive 

Summary). This sequence represents the process by which we can determine the risks associated 

with ambient concentrations of NOx and SOx. However, for the purposes of discussing a 

conceptual framework for design of standards to protect against those risks, a modified version 

of the risk framework is needed. 

Figure 1-1 depicts the framework by which we are considering the structure of an 

ecologically meaningful secondary standard. It is a conceptual diagram that illustrates how a 

level of protection related to an indicator of ecological effect(s) equates to atmospheric 

concentrations of NOx and SOx indicators. This conceptual diagram illustrates the linkages 

between ambient air concentrations and resulting deposition metrics, and between the deposition 

metric and the ecological indicator of concern. The Atmospheric Deposition Transformation 

Function translates ambient atmospheric concentrations of NOx and SOx to nitrogen and sulfur 

deposition metrics, while the Ecological Effect Function transforms the deposition metric into 

the ecological indicator. 

Development of a form for the standard that reflects this structure is a critical step in the 

overall standard setting process. The atmospheric levels of NOx and SOx that satisfy 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 of the ecological indicator for a targeted effect. 

The details of this conceptual framework are discussed in Chapter 5, including 

discussions of modifying factors that alter the relationship between ambient atmospheric 

concentrations of NOx and SOx and depositional loads of nitrogen and sulfur, and those that 

modify the relationship between deposition loads and the ecological indicator. 

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In setting NAAQS to protect public health and welfare, EPA has historically established 

standards which require the comparison of monitored concentrations of an air pollutant against a 

numerical metric of atmospheric concentration that does not vary geographically. This approach 

has appropriately protected public health as at-risk populations are widely distributed throughout 

the nation. As more is learned about the effects of pollutants such as NOx and SOx and the 

environment, however, such an approach may not be appropriate to provide the requisite level of 

protection to public welfare from effects on sensitive ecosystems. EPA is proposing in this 

review of the secondary standard for NOx and SOx a standard that takes into account variable 

factors, such as atmospheric variables and location-specific characteristics of ecosystems, as the 

appropriate approach to protect the public welfare from the effects associated with the presence 

of these pollutants in the ambient air. 

EPA must undertake a thorough review of the air quality criteria for the pollutant at issue 

in reviewing a secondary NAAQS, and determine whether a current standard is requisite to 

protect the public welfare. Under section 108 of the CAA, air quality criteria are to “reflect the 

latest scientific knowledge useful in indicating the kind and extent of all identifiable effects” 

associated with the presence of the pollutant in the ambient air. It is clear from the language of 

the CAA that where the state of the science provides a basis for considering such effects, the 

review of the air quality criteria should encompass a broad analysis of “any” known or 

anticipated adverse effects, as well as the ways in which variable conditions such as atmospheric 

conditions may impact the effect of a pollutant and the ways in which other air pollutants may 

interact with the criteria pollutant to produce adverse effects. Specifically, section 108(a)(2) of 

the CAA provides that: 

Air quality criteria for an air pollutant shall 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 such pollutant in the ambient air, in varying 

quantities. The criteria for an air pollutant to the extent practicable, shall include information on: 

(A) Those variable factors (including atmospheric conditions) which of themselves or 

in combination with other factors may alter the effects on public health or welfare 

of such air pollutants; 

(B) The types of air pollutants which, when present in the atmosphere, may interact 

with such pollutants to produce an adverse effect on public health or welfare; and 

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(C) Any known or anticipated adverse effects on welfare. 

 Based on this extensive review of the air quality criteria for an air pollutant, the 

Administrator is required to review and to revise, as appropriate, the secondary standard to 

ensure that the standard “is requisite to protect public welfare from any known or anticipated 

adverse effects associated with the presence of such air pollutant in the ambient air.” CAA § 

109(b) & (d). “Effects on welfare,” in turn, is defined to include a broad array of effects, 

including effects on soil, water, crops, vegetation, and manmade materials, “whether caused by 

transformation, conversion, or combination with other air pollutants.” CAA § 302(h). Thus, as 

with the sections of the CAA describing the issuance of air quality criteria, the CAA uses 

expansive language in describing the scope of EPA’s responsibility and the range of effects that 

EPA should take into account in setting a standard that is requisite to protect public welfare. The 

term “requisite,” however, indicates that section 109 is not open-ended. In considering the 

meaning of the term “requisite” in the context of the primary standards, the Supreme Court has 

agreed with EPA that such a standard is one that is “sufficient, but not more than necessary” to 

protect public health. Whitman v. American Trucking, 531 U.S. 457, 473 (2001). 

While EPA has most often considered the results of direct exposure to an air pollutant in 

the ambient air in assessing effects on public health and welfare, such as the health effects on 

humans when breathing in an air pollutant or the effects on vegetation through the uptake of air 

pollutants from the ambient air through leaves, EPA has also considered, where appropriate, the 

effects of exposure to air pollutants through more indirect mechanisms. For example, both in 

1978 and in 2008, EPA established a NAAQS for lead that addressed the health effects of 

ambient lead whether the lead particles were inhaled or were ingested after deposition on the 

ground or other surfaces. 73 FR 66964 (November 12, 2008), Lead Industries v. EPA, 647 F.2d 

1130 (DC Cir. 1980) (1978 NAAQS). The deposition of ambient NOx and SOx to terrestrial and 

aquatic environments can impact ecosystems through both direct and indirect mechanisms, as 

discussed in the REA and this document. Given Congress’ instruction to set a standard that “is 

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

with the presence of such air pollutant in the ambient air,” 42 U.S.C. § 109 (b)(2), this review 

appropriately attempts to take into consideration widely acknowledged effects, such as 

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acidification and nutrient enrichment, which are associated with the presence of ambient SOx 

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In this review, EPA is also attempting to develop a standard that takes into account the 

variability in effects from ambient levels of SOx and NOx. The CAA requires EPA to establish 

“national” standards, based on the air quality criteria that provide the requisite degree of 

protection, but does not clearly address how to do so under the circumstances present here. One 

approach is to develop a secondary standard such as the one discussed in this Policy Assessment 

Document. Such a standard is designed to provide a generally uniform degree of protection 

throughout the country by allowing for varying concentrations of allowable ambient NOx and 

SOx, depending on atmospheric conditions and other variabilities, to achieve that degree of 

protection5

. Such a standard protects sensitive ecosystems wherever such ecosystems are found. 

This approach recognizes that setting a standard that is sufficient to protect the public welfare but 

not more than is necessary calls for consideration of a standard such as the one discussed in this 

document. 

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 In concept, this approach to setting a national standard using a consistent form that may result in differing levels of 

atmospheric concentrations fo NOx and SOx is similar to the current proposal to use PM10 as the national indicator 

for protection against the health effects of PM10-2.5, as discussed in the 2010 2nd draft Policy Assessment for the 

PM NAAQS primary standards. In that case, EPA is proposing to recognize that the same size fraction of particles 

may have different toxicity depending on location, and as a result, it is more appropriate to use an indicator that 

reflects that varying toxicity, rather than setting one absolute level of PM10-2.5 which may not be equally protective 

in all locations. By proposing this form, EPA is recognizing that in attaining a PM10 standard, the resulting balance 

of PM10-2.5 and PM2.5 particles will differ across areas of the U.S. Likewise, setting a joint NOx/SOx standard 

that results in differing allowable concentrations of NOx and SOx across the U.S. based on the differing potential for 

NOx and SOx to result in ecological damages is appropriate in providing a requisite level of welfare protection. 

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Structure of an Ecologically-based Standard

Air Quality 

Indicator(s)

Measured over a 

specified averaging 

time; expressed in 

terms of a specified statistic (form)

Ecological 

Indicator

Ecological 

Effect

Function

Atmospheric 

Deposition

Transformation

Function

Deposition 

Metric

Variable/Fixed 

Factors:

Ecological

Variable/Fixed 

Factors:

Atmospheric

Landscape

Form of the Standard

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Figure 1-1. Framework of an alternative secondary standard. 

 1.6 POLICY RELEVANT QUESTIONS 

In this policy assessment, a series of general questions frames our approach to identifying 

a range of policy options for consideration by the Administrator regarding secondary NAAQS 

for NOx and SOx. These questions are drawn from our Integrated Review Plan with 

modifications based on further consideration by staff and comments from CASAC and the 

public. Our policy assessment begins by characterizing “known or anticipated adverse effects” 

on public welfare within our conceptual model (CPE 1). As noted earlier, this review is focusing 

on effects in sensitive unmanaged ecosystems (not commercial forests or agricultural lands6

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resulting from ambient concentrations of NOx and SOx through deposition of N and S. 

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 The decision to focus on unmanaged ecosystems is based on the weight of evidence of effects in those ecosystems. 

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 N deposition 

resulting from atmospheric N (ISA 3.3.2.5). In addition ,the ISA notes that agricultural and commercial forest lands 

are routinely fertilized with amounts of N (100 to 300 kg N/ha) that exceed air pollutant inputs even in the most 

polluted areas (ISA 3.3.9). This review recognizes that effect of N deposition in managed areas may be viewed 

differently than effects of N deposition in 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. 

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 In Chapter 2, we draw from the information and conclusions presented in the ISA and 

REA to address the following questions: 

1. What are the nature and magnitude of ecosystem responses to reactive nitrogen and 

sulfur deposition, acidification, nutrient depletion and the mobilization of toxic metals 

in sensitive aquatic and terrestrial ecosystems? 

a. How are these responses affected by landscape factors? 

b. What types of ecosystems are sensitive to such responses? 

2. To what extent can ecosystem responses to nitrogen deposition be separated into 

responses related to oxidized and reduced forms of reactive nitrogen compounds? 

In Chapter 3, we address the following questions related to linking effects to measures of 

adversity (CPE 1.1): 

1. How do we characterize adversity to public welfare? What are the sources of 

potentially relevant characterization for this policy assessment? 

2. What is the evidence of effects on ecosystem services, and how can those ecosystem 

services be linked to ecological indicators? 

3. To what extent are identified ecosystem effects important from a public welfare 

perspective, and what are the important uncertainties associated with estimating such 

effects? 

Once we have described ecological effects, we then provide an assessment of the 

adequacy of the existing NOx and SOx standards (CPE 1.2). We begin this assessment by 

drawing from the information and conclusions presented in the ISA and REA to address in 

Chapter 4 the following questions, which allow us to identify whether the structure of the current 

standards is appropriate relative to the key ecological effects assessed in the ISA and REA, 

including acidification and excess nutrient enrichment and whether there is adequate information 

and analyses available at this time to assess the extent to which potentially adverse effects on 

aquatic and terrestrial ecosystems can be associated with current levels of atmospheric reactive 

nitrogen, accounting for the contributions of oxidized and reduced forms, and SOx and with 

levels that are at or below the current secondary standards: 

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1. To what extent are effects that could reasonably be judged to be adverse to public 

welfare occurring under current conditions and would such effects occur if the nation 

met the current standards? To what extent do the current NOx and SOx secondary 

standards provide protection from effects associated with deposition of: 

a. Sulfur and oxidized nitrogen from atmospheric NOx, and SOx which results in 

acidification in sensitive aquatic and terrestrial ecosystems? 

b. Oxidized nitrogen from atmospheric NOx, which results in nutrient enrichment 

effects in sensitive aquatic and terrestrial ecosystems? 

c. Sulfur and oxidized nitrogen from atmospheric NOx and SOx which results in 

other ecological effects (e.g. mercury methylation)? 

2. In what way are the structures of the current NOx and SOx secondary standards 

inadequate to protect against public welfare effects? 

In Chapter 5, we follow our adequacy assessment by developing in greater detail the 

conceptual framework for the design of ecologically relevant multi-pollutant standards 

introduced in Section 1.4 above. To the extent that the available information calls into question 

the adequacy of protection afforded by the current standards and/or the appropriateness of the 

structure of the standards, we explore the extent to which available information supports 

consideration of alternative standards, in terms of atmospheric and ecological indicators and 

related averaging times, forms, and levels. This conceptual framework is designed to focus on 

resolving the following questions: 

1. (CPE 2.1) Does the available information provide support for the use of ecological 

indicators to characterize the responses of aquatic and terrestrial ecosystems to 

oxidized nitrogen and sulfur deposition? 

2. (CPE 1) Does the available information provide support for the development of 

appropriate ecological response to deposition relationship(s) that meaningfully relates 

oxidized nitrogen and sulfur deposition to relevant ecological indicators? Does a 

quantified relationship exist between the level of a relevant ecological indicator and 

an amount of nitrogen and sulfur deposition? 

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3. (CPE 2) What are the important variables in the ecological response to deposition 

relationship(s)? Are these relationships applicable nationally? What are the 

appropriate temporal scales for these relationships? 

a. How does ecological response to deposition relationship(s) depend upon 

spatially heterogeneous geologic factors (e.g. bedrock type, weathering rates) 

that govern sensitivity? 

b. How do we consider areas with high natural background acidification or 

nutrient loadings? 

4. (CPE 3) Does the available information provide support for the development of 

appropriate functions that characterize the relationships between atmospheric NOx 

and SOx and the wet and dry deposition of total reactive nitrogen and sulfur? 

a. What deposition function is appropriate to use for the purpose of relating an 

amount of nitrogen and/or sulfur deposition in sensitive ecosystems to 

ambient concentrations of atmospheric reactive nitrogen, including oxides and 

reduced forms, and/or sulfur? What are the important variables in such a 

function? What are appropriate spatial and temporal scales to use in 

specifying such variables? 

Based on the conceptual framework for the structure of the ecologically relevant multipollutant standards, we then address in Chapter 6 the components of the standard needed to 

develop options for consideration by the Administrator. Development of these options will focus 

on addressing the following questions: 

1. (CPE 2.1) What ecological indicators are appropriate to use for the purpose of 

developing an alternative standard for the various ecological effects assessed in this 

review? 

2. (CPE 5) What indicators of oxides of nitrogen and sulfur are appropriate to use for 

the purpose of determining whether the resultant deposition is within the target values 

needed to achieve the desired degree of protection? What averaging times and forms 

are appropriate to consider? 

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3. (CPE 2) What approaches are available to specify non-atmospheric components of 

the standard, e.g. weathering rates? Are there approaches that can simplify the 

structure of the standard? 

4. What are the available approaches for accounting for reduced N in the structure of the 

standard? 

5. What is the most appropriate form for the standards to reflect the relationships 

between ambient NOx and SOx, acidifying deposition, and the ecological indicator 

for acidification? 

Several follow-up questions derive from our assessment of options for specifying the 

components of a multipollutant standard. In Chapter 6, we address the questions: 

1. To what extent would a standard specifically defined to protect against one ecological 

effect (i.e., aquatic acidification) likely provide protection from other relevant 

ecological effects? 

2. What are the available approaches for combining multiple indicators into a single 

standard, e.g. using nitrogen effects to bound the tradeoff curve for NOx/SOx for 

aquatic acidification effects 

3. What are the available approaches to integrate potential standards for aquatic and 

terrestrial acidification and/or aquatic and terrestrial N enrichment? 

In Chapter 7, we provide a range of explorations of uncertainties in the evidence and 

models as they pertain to the selection of options for components of the standard. In addition, 

we provide results of sensitivity analyses for components of the proposed AAPI form, as well as 

characterizing information on variability in those components. The chapter focuses on the 

following questions: 

1. What are critical uncertainties in the characterization of pre-industrial levels of 

ANC? 

2. What uncertainties are introduced through the use of steady state critical load 

models relative to dynamic critical loads models? 

3. What are the critical uncertanties in the modeled relationship between 

concentrations of NOx and SOx and deposition of N and S? 

4. What are the critical uncertainties in the modeled values of NHx deposition? 

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5. How sensitive is the form of the NOx/SOx standard to the primary components? 

6. How sensitive are the atmospheric transformation ratios calculated from CMAQ 

to the emissions scenarios, chemical mechanisms, and meteorology in the model? 

7. How well does CMAQ perform in simulating nitric oxide, nitrogen dioxide, sulfur 

dioxide, nitrate, ammonium and aerosol nitrate, ammonium, and sulfate relative to 

observations from different networks for which the data are routinely available? 

8. How well does CMAQ perform in simulating wet deposition of sulfate, nitrate, 

and ammonium relative to observations from the National Atmospheric 

Deposition Program (NADP) network? 

Chapter 8 provides a discussion of several important aspects of monitoring of NOx and 

SOx, including methods, network design, and frequency. 

We conclude in Chapter 9 with a discussion of options to consider in selecting pollutant 

indicators, averaging times, forms, and ranges of levels for the secondary NOx and SOx 

standards. This discussion is informed by a consideration of the role of ecosystem services in 

helping to characterize what adversity to public welfare, focused on the following questions: 

1. (CPE 5) What are the risks of ecosystem service impairment under alternative levels 

of potential standards for NOx and SOx? 

2. (CPE 5) To what extent can information about ecosystem services be used to help 

characterize the extent to which differing levels of relevant ecological indicators 

reflect impacts that can reasonably be judged to be adverse from a public welfare 

perspective? 

3. (CPE 5) Are there relevant benchmarks for adversity to public welfare that can be 

derived from other sources? 

4. (CPE 5) Taking into consideration information about ecosystem services and other 

factors related to characterizing adversity to public welfare for the ecological effects 

being assessed in this review, what is an appropriate range of levels of protection to 

be achieved by alternative standards for the Agency to consider? 
1 2 KNOWN OR ANTICIPATED ECOLOGICAL EFFECTS 

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In this chapter we address Critical Policy Element 1, evaluation of the effects of ambient 

NOx and SOx on ecosystems, and the relationship between those effects and the measure of dose 

in the ecosystem, indicated by the depositional loadings of N and S. In section 302(h) of the 

Clean Air Act, welfare effects addressed by a secondary NAAQS 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”. Of these welfare 

effects categories, the effects of NOx and SOx on aquatic and terrestrial ecosystems, which 

encompass soils, water, vegetation, wildlife, and contribute to economic value and well-being, 

are of most concern at concentrations typically occurring in the U.S. Direct effects of NOx and 

SOx on vegetation are also discussed in this chapter, and have been the focus of previous 

reviews. However, for this review, the focus of this chapter is on the known and anticipated 

effects to ecosystems caused by exposure to NOx and SOx through deposition. 

The information presented here is a concise summary of conclusions from the ISA and 

the REA. This chapter focuses on effects on specific ecosystems with a brief discussion on 

critical uncertainties associated with acidification and nutrient enrichment. Those effects are then 

evaluated in Chapter 3 within the context of alternative definitions of, including assessments of 

potential impacts on ecosystem services. Effects are broadly categorized into acidification and 

nutrient-enrichment in the proceeding sections. This is background information intended to 

support new approaches for the design of ecologically relevant secondary NOx and SOx 

standards which are protective of U.S. ecosystems. More detailed information on the conceptual 

design and specific options for the proposed standards are presented in Chapters 5 and 9of this 

policy assessment document. While we provide a summary of effects for all four of the primary 

effects categories, we reiterate that the focus of this second draft policy assessment is on effects 

related to aquatic acidification, without downplaying the potential importance of effects in other 

categories. 

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2.1 ACIDIFICATION: EVIDENCE OF EFFECTS ON STRUCTURE AND 

FUNCTION OF TERRESTRIAL AND FRESHWATER ECOSYSTEMS 

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Sulfur oxides (SOx) and nitrogen oxides (NOx) 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 buffered by high base soils, 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 waterbodies (REA 2.1). The chemical forms of nitrogen that 

may contribute to acidifying deposition include both oxidized and reduced chemical species. 

When sulfur or nitrogen (NOx, NHx and Nr) leaches from soils to surface waters in the 

form of SO4

2- or NO3-, an equivalent amount of positive cations, or countercharge, is also 

transported. This maintains electroneutrality. If the countercharge is provided by base cations, 

such as calcium (Ca2+), magnesium (Mg2+), sodium (Na+), or potassium (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. 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. 

(REA 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. For the purpose of a national secondary standard, the most relevant characteristics are 

those that are less variable over small scales. 

Acidifying deposition of NOx and SOx 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 N and S to ecosystems and the 

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

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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 was passing 

through deeper soil horizons (REA 4.2). 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 United States. Acidification affects both terrestrial and 

freshwater aquatic ecosystems. Terrestrial and aquatic processes are often linked; therefore 

responses to the following questions address both types of ecosystems unless otherwise noted. 

2.1.1 What is the nature of acidification related ecosystem responses to reactive 

nitrogen and sulfur deposition? 

The ISA concluded that deposition of SOx, NOx, and NHx leads to the varying degrees of 

acidification of ecosystems (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. 

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

deposition and effects on biogeochemistry and biota in aquatic ecosystems (ISA 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. (ISA 3.2.3.2). 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 acidsensitive biological species (ISA 3.2.3.3). 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 (Baker et al., 1990b). Biological effects are linked to changes in water chemistry 

including decreases in ANC and pH and increases in inorganic Al concentration. The direct 

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biological effects are caused by lowered pH and increased inorganic Al concentrations. While 

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

acidification (Fig 2-1, See further discussion in Section 2.1.3). 

↑ Al+3

↑Ecological effects which may include: 

Lower biodiversity 

Altered species composition 

Localized extinction of sensitive species 

Individual mortality of sensitive species 

Sub-lethal effects to sensitive species and 

ecological integrity 

 pH 

Stream water chemistry 

Aquatic biota 

 ANC

↑ NO3- and SO4

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Figure 2-1. Conceptual model of direct and indirect acidification effects on aquatic biota. 

Acidic pollutants (NO3- and SO4-2) lower ANC, resulting in lower pH with 

direct toxic effects on fish. The lower pH mobilizes Al3+ from soils often 

resulting in higher concentration in stream water causing direct toxicity to 

fish. 

These changes in stream water chemistry contribute to declines in taxonomic richness of 

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

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

recreation for humans (see Chapter 3). 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 (ISA Section 4.2.2.2. and Table 3-9). For example, breeding 

distribution of the common goldeneye (Bucephala clangula) an insectivorous duck, may be 

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affected by changes in acidifying deposition (Longcore and Gill, 1993). 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 (Parker 1988). 

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

acidifying deposition and changes in biogeochemistry (ISA 4.2.1.1). The strongest evidence 

comes from studies of forested ecosystems, with supportive information on other plant taxa, 

including shrubs and lichens (ISA 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 C:N ratio (ISA 

3.2.2.2). 

In soils with base saturation less than about 15 to 20%, exchange chemistry is dominated 

by Al (Reuss, 1983). Under these conditions, responses to inputs of sulfuric acid and nitric acid 

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 (Parker et al., 1989; U.S. EPA, 2009). 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 (Aber et al., 2003; Ross et al., 2004). 

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 (DeHayes et al., 1999). The frequency of freezing injury to red 

spruce needles has increased over the past 40 years, a period that coincided with increased 

emissions of S and N oxides and increased acidifying deposition (DeHayes et al., 1999). 

Acidifying deposition can contribute to dieback in sugar maple (Acer saccharum) through 

depletion of cations from soil with low levels of available Ca (Horsley et al., 2000; Bailey et al., 

2004). Grasslands are likely less sensitive to acidification than forests due to grassland soils 

being generally rich in base cations (Fenn et al., 2003; Blake et al., 1999). 

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2.1.2 What types of ecosystems are sensitive to such effects? In which ways are these 

responses affected by atmospheric, ecological, and landscape factors? 

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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 (ISA 3.2.4.1). 

2.1.3 What is the magnitude of ecosystem responses to acidifying deposition? 

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

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

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

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

2.1.3.1 Aquatic 

Freshwater ecosystem surveys and monitoring in the eastern United States 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) (Stoddard, 1990), and Long-term Monitoring 

(LTM) (Ford et al., 1993; Stoddard et al., 1996) 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, 

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acidification of surface waters by acidic deposition is uncommon 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 (ISA 3.2.4.2 and REA 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 (ISA 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 (Keller 

and Gunn 1995; Schindler et al., 1985). 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) (Baker et al., 1990). 

In response to small to moderate changes in acidity, acid-sensitive species are often replaced by 

other more acid-tolerant species, resulting in changes in community composition and richness. 

The effects of acidification are continuous, with more species being affected at higher degrees of 

acidification. At a point, typically a pH <4.5 and an ANC <0 μeq/L, complete to near-complete 

loss of many taxa of organisms occur, including fish and aquatic insect populations, whereas 

other taxa are reduced to only acidophilic species. These changes in taxa composition are 

associated with the high energy cost in maintaining physiological homeostasis, growth, and 

reproduction at low ANC levels (Schreck, 1981, 1982; Wedemeger et al., 1990; REA appendix 

2.3). Decreases in species richness related to acidification have been observed in the Adirondack 

Mountains and Catskill Mountains of New York (Baker et al., 1996), New England and 

Pennsylvania (Haines and Baker, 1986), and Virginia (Bulger et al., 2000). 

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 (REA Chapter 4). 

 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; ISA). 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. 

ANC values were categorized according to their effects on biota, as shown in Figure 2-2. [Need 

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to indicate the source for the categorization] 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. 

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

 neutralizing capacity (ANC) 

 

The most commonly used models of acidification are presented in Table 2-2 These models are 

designed to be applied at the spatial scale of the watershed, with the exception of the SMART 

model. Steady-sate mass balance models, including the steady-state water chemistry (SSWC) 

model are the most commonly used method for analysis of critical loads of acid deposition. The 

steady-state models assume steady state conditions. The dynamic models consider how the 

ecosystem may change through time. These models tend to require more data than the steadystate models. 

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Table 2-2 Summary several commonly used acidification models (See ISA Annex 

A for a more comprehensive list and discussion of acidification models) 

Model name Dynamic 

or steady 

state 

Model description 

Steady-sate mass 

balance models 

/Steady-state 

water Chemistry 

(SSWC)/ 

Steadystate The basic principle is based on identifying the long-term 

average sources of acidity and alkalinity in order to 

determine the maximum acid input that will balance the 

system at a biogeochemical safe-limit. Several 

assumptions have been made in the steady state 

calculations. First, it is assumed that ion exchange is at 

steady state and there is no net change in basesaturation or 

no net transfer of ANC from soil solution to the ion 

exchange matrix. It is assumed that for N there is no net 

denitrification, adsorption or desorption and the N cycle is 

at steady state. Sulfate is also assumed to be at steady 

state: no sulfide oxidation, sulfate uptake, sulfate 

permanent fixation or sulfate reduction are significant. 

Simple hydrology is assumed where there is straight 

infiltration through the soil profile. 

Model of 

Acidification of 

Groundwater in 

Catchment 

(MAGIC) 

Dynamic MAGIC is a lumped-parameter model of intermediate 

complexity, developed to predict the longterm effects of 

acidic deposition on surface water chemistry. The model 

simulates soil solution chemistry and surface water 

chemistry to predict the monthly and annual average 

concentrations of the major ions in these waters. MAGIC 

consists of: a section in which the concentrations of major 

ions 

are assumed to be governed by simultaneous reactions 

involving SO4 2− adsorption, cation exchange, 

dissolution-precipitation- speciation of aluminum, and 

dissolution-speciation of inorganic carbon; and a 

mass balance section in which the flux of major ions to 

and from the soil is assumed to be controlled by 

atmospheric inputs, chemical weathering, net uptake and 

loss in biomass and losses to runoff. 

PnET-BGC Dynamic PnET/BGC simulates major biogeochemical processes, 

such as forest canopy element transformations, hydrology, 

soil organic matter dynamics, N cycling, geochemical 

weathering, and chemical equilibrium reactions in solid 

and solution phases, and allows for simulations of land 

disturbance. The model uses mass transfer relationships to 

describe weathering, canopy interactions and surface 

water processes. Chemical equilibrium relationships 

describe anion adsorption, cation exchange and soil 

solution and surface water speciation. The model operates 

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Table 2-2 Summary several commonly used acidification models (See ISA Annex 

A for a more comprehensive list and discussion of acidification models) 

on a monthly time step and is applied at the stand to 

small-watershed scale. 

DayCent-Chem Dynamic DayCent-Chem links two widely accepted and tested 

models, one of daily biogeochemistry for forest, grassland, 

cropland, and savanna systems, DayCent (Parton et al., 

1998), and the other of soil and water geochemical 

equilibrium, PHREEQC (Parkhurst and Appelo, 1999). 

The linked DayCent/PHREEQC model was created to 

capture the biogeochemical responses to atmospheric 

deposition and to explicitly consider those biogeochemical 

influences on soil and surface water chemistry. The linked 

model expands on DayCent’s ability to simulate N, P, S, 

and C ecosystem dynamics byincorporating the reactions 

of many other chemical species in surface water. 

Very Simple 

Dynamic (VSD) 

soil acidification 

model 

Dynamic This model is frequently used in Europe to simulate 

acidification effects in soils when observed data are 

sparse. The VSD model consists of a set of mass balance 

equations, describing the soil input-output relationships, 

and a set of equations describing the rate-limited and 

equilibrium soil processes.It only includes weathering, 

cation exchange, N immobilization processes, and a mass 

balance for cations, sulfur and N. In the VSD model, the 

various ecosystem processes have been limited to a few 

key processes. Processes that are not taken into account 

include canopy interactions; nutrient cycling processes; N 

fixation and NH4 adsorption; SO42− transformations 

(adsorption, uptake, immobilization, and reduction); 

formation and protonation of organic anions; and 

complexation of Al. 

Simulation 

Model for 

Acidification’s 

Regional Trends 

(SMART) 

Dynamic The the SMART model consists of a set of mass balance 

equations, describing soil input/output relationships, and a 

set of equations describing the rate-limited and 

equilibrium soil processes. It 

includes most of the assumptions and simplifications 

given for the VSD model. SMART models the exchange 

of Al, H, and divalent base cations using Gaines Thomas 

equations. Additionally, SO4 

2− adsorption is modeled using a Langmuir equation (as 

in MAGIC) and organic acids can be described as mono-, 

di-, or tri-protic. The SMART model has been developed 

with regional applications in mind, and an early example 

of an application to Europe can be found in De Vries et al. 

(1994). 

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The analyses of the Adirondack Case Study Area indicated that although wet deposition 

rates for SO2 and NOx have been reduced since the mid-1990s, current concentrations are still 

well above simulated pre-acidification (1860) conditions. Modeling predicts NO3

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 and SO4

2- are 

17- and 5-fold higher, respectively, in 2006 than under simulated pre-acidification conditions. 

Based on the 2006 Model of Acidification of Groundwater in Catchment (MAGIC) simulations, 

the estimated average ANC across the 44 lakes in the Adirondack Case Study Area is 62.1 μeq/L 

(± 15.7 μeq/L); 78 % of all monitored lakes in the Adirondack Case Study Area have a current 

risk of Elevated, Severe, or Acute. Of the 78%, 18% are chronically acidic (REA 4.2.4.2). 

Based on a N and S deposition scenario that maintains current emission levels to 2020 

and 2050, the simulation forecast indicates there would be 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 from 2020 to 2050. 

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Figure 2-2. Average NO3

- concentrations (orange), SO42- concentrations 

 (red), and ANC (blue) across the 44 lakes in the Adirondack 

 Case Study Area modeled using MAGIC for the period 1850 

 to 2050. 

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Figure 2-3. ANC concentrations of preacidification (1860) and 2006 conditions 

based on hindcasts of 44 lakes in the Adirondack Case Study Area 

modeled using MAGIC. 

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Figure 2-4. Critical loads of acidifying deposition that each surface water location 

can receive in the Adirondack Case Study Area while maintaining or 

exceeding an ANC concentration of 50 μeq/L based on 2002 data. 

Watersheds with critical load values <100 meq/m2

/yr (red and orange 

circles) are most sensitive to surface water acidification, whereas 

watersheds with values >100 meq/m2/yr (yellow and green circles) are 

less sensitive sites. 

Note that 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) (Kretser et al., 1989), 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 

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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 (Driscoll et al., 2001; Kretser et al., 1989). In the Adirondacks, a positive relationship 

exists between the pH and ANC in lakes and the number of fish species present in those lakes 

(ISA 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. MAGIC modeling predicts surface water 

concentrations of NO3- and SO4

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

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

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

risk of Elevated, Severe, or Acute. Of the 55%, 18% are chronically acidic today (REA 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 still have 

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

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

Moderate Concern decreases by 5%. 

Biological effects of increased acidification documented in the Shenandoah Case Study 

Area include a decrease in the condition factor in blacknose dace (Dennis and Bulgar 1995, 

Bulgar et al., 1999) and a decrease in fish biodiversity associated with decreasing stream ANC 

(Bulger et al.,1995; Dennis and Bulger, 1995; Dennis et al., 1995; MacAvoy and Bulger, 1995, 

Bulgar et al., 1999). 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 (ISA 3.2.3.4). 

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Figure 2-5. Average NO3

- concentrations orange), SO42-concentrations (red), and 

ANC (blue) levels for the 60 streams in the Shenandoah Case Study 

Area modeled using MAGIC for the period 1850 to 2050. 

 

Figure 2-6. ANC levels of 1860 (preacidification) and 2006 (current) conditions 

based on hindcasts of 60 streams in the Shenandoah Case Study Area 

modeled using MAGIC. 

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Figure 2-7. Critical loads of surface water acidity for an ANC of 50 μeq/L for 

Shenandoah Case Study Area streams. Each dot represents an 

estimated amount of acidifying deposition (i.e., critical load) that each 

stream’s watershed can receive and still maintain a surface water 

ANC >50 μeq/L. Watersheds with critical load values <100 meq/m2/yr 

(red and orange circles) are most sensitive to surface water 

acidification, whereas watersheds with values >100 meq/m2/yr (yellow 

and green circles) are less sensitive sites. 

2.1.3.2 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 potassium), as well as soil 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 

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variety of tree species ranged from 0.2 to 0.8 (Sverdrup and Warfvinge, 1993, a meta-data 

analysis of laboratory and field studies). 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 (Cronan and Grigal, 1995; meta-data assessment) (REA pg 4-54, REA Appendix 5). 

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% of forest ecosystems in the 

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

>250 eq ha-1

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-1 (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 (ISA 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 United States 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%. Sugar maple growth can be decreased by 20% at a 

Bc/Al ratio of 0.6 (REA 4.4). The REA analysis determined the health of at least a portion of the 

sugar maple and red spruce growing in the United States 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% to 75% of all sugar maple 

plots across 24 states. For red spruce, total nitrogen and sulfur deposition levels exceeded three 

selected critical loads in 3% to 36% of all red spruce plots across eight states (REA 4.4). 

2.1.4 What are the 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, 

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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 (Figure 1-1). 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 (REA 7.0): 

6  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. 

 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. 

 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. 

2.1.4.1 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 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 was derived from multiple 

site calibrations. The 95% confidence interval for pre-acidification of lakes was an average of 15 

eq/L difference in ANC concentrations or 10% and 8 eq/L or 5% for streams (REA 7.1.2). 

The use of the critical load model used 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., 

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2007). To propagate uncertainty in the model parameters, Monte Carlo methods were employed 

to develop an inverse function of exceedences. There is high confidence associated with the 

ecological effect function developed for aquatic acidification. In calculating the ANC function, 

the depositional load for N or S is fixed by the deposition of the other, so deposition for either 

will never be zero (Figure 7.1-6 REA). 

2.1.4.2 Terrestrial Acidification 

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

robust and considered high quality. The USFS-Kane Experimental Forest and significant 

amounts of research work in the Allegheny Plateau have produced extensive, peer-reviewed data 

sets. A meta-analysis of laboratory studies showed that tree growth was decreased by 20% 

relative to controls for BC/Al ratios (ISA 7.2.1 and Figure 7.2-1). Sugar maple and red spruce 

were the focus of the REA since they are demonstrated to be negatively affected by soil available 

Ca2+ 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 (ISA 7.2.2). There is fairly high confidence associated with the ecological 

effect function developed for terrestrial acidification (REA 7.2.3). 

2.2 NITROGEN ENRICHMENT: EVIDENCE OF EFFECTS ON STRUCTURE AND 

FUNCTION OF TERRESTRIAL AND FRESHWATER ECOSYSTEMS 

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 N greatly exceed N deposition. Both oxides of nitrogen and reduced forms of 

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nitrogen, (e.g. 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. We address the issue of 

incorporating the relative contributions of oxidized and reduced nitrogen into the standards in 

Chapters 5 and 8. 

2.2.1 What is the nature of terrestrial and freshwater ecosystem responses to reactive 

nitrogen and/ sulfur deposition? 

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

enrichment of ecosystems (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. 

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

alteration of biogeochemical cycling in terrestrial ecosystems (ISA 4.3.1.1 and 3.3.2.1). This is 

supported by numerous observational, deposition gradient and field addition experiments in 

sensitive ecosystems. Stoddard (1994) identified the leaching of NO3- in soil drainage waters and 

the export of NO3- in steam water as two of the primary indictors of N enrichment. Several Naddition studies indicate that NO3- leaching is induced by chronic additions of N (Edwards et al., 

2002b; Kahl et al., 1999; Peterjohn et al., 1996; Norton et al., 1999). Aber et al. (2003) found 

that surface water NO3- concentrations exceeded eq/L in watersheds receiving about 9 to 13 

kg N/ha/yr of atmospheric N deposition. N deposition disrupts the nutrient balance of 

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

measured include NO3

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

denitrification, foliar N concentration, and soil water NO3 − and NH4+ concentrations. Note that 

N saturation (N 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 N and C cycling, the effects of N on C budgets (quantified input and 

output of C 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 N deposition has been shown to cause increased litter 

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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 N deposition 

on C budgets of non-forest ecosystems. 

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

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

(ISA 4.3.1.2). Some organisms and ecosystems are more sensitive to N deposition and effects of 

N deposition are not observed in all habitats. The most sensitive terrestrial taxa to N 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 N 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 nonnative, invasive species are associated with N deposition (Clark and Tillman, 2008; Schwinning 

et al., 2005). 

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

N deposition and the alteration of biogeochemical cycling in freshwater aquatic ecosystems (ISA 

3.3.2.3). N deposition is the main source of N enrichment to headwater streams, lower order 

streams and high elevation lakes. The most common chemical indicators that were studied 

included NO3

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Chl a:total P ratio. Elevated surface water NO3

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western U.S. Bergstrom and Jansson (2006) report a significant correlation between N deposition 

and lake biogeochemistry by identifying a correlation between wet deposition and [DIN] and Chl 

a: Total P. Recent evidence provides examples of lakes and streams that are limited by N and 

show signs of eutrophication in response to N addition. 

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

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

ecosystems (ISA 3.3.5.3). Increased N deposition can cause a shift in community composition 

and reduce algal biodiversity, especially in sensitive oligotrophic lakes. 

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2.2.2 What types of ecosystems are sensitive to such effects? How are these responses 

affected by atmospheric, ecological, and landscape factors? 

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

sensitivity to N deposition (Clark and Tilman 2008; Aber et al., 2003; Fenn et al., 2003; Rueth et 

al., 2003; Egerton-Warburton and Allen 2000; Williams et al., 1996; and additional studies 

summarized in Table 4-4 ISA). Increased deposition to N-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 N, may result 

in imbalance in ecological stoichiometry, with effects on ecosystem processes, structure and 

function (Sterner and Elser, 2002). In general, N 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 N deposition levels as low as 3 kg N/ha/yr in 

sensitive terrestrial ecosystems to N deposition. In aquatic ecosystems, N 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 N loading and N-sensitivity. 

N-sensitivity is predominately driven by the degree to which growth is limited by nitrogen 

availability. Grasslands in the western United States are typically N-limited ecosystems 

dominated by a diverse mix of perennial forbs and grass species (Clark and Tilman, 2008; 

Suding et al., 2005). A meta-analysis by Lebauer and Treseder (2008) indicated that N 

fertilization increased aboveground growth in all non-forest ecosystems except for deserts. In 

other words, almost all terrestrial ecosystems are N-limited and will be altered by the addition of 

anthropogenic nitrogen (LeBauer and Treseder, 2008). Likewise, a freshwater lake or stream 

must be N-limited to be sensitive to N-mediated eutrophication. There are many examples of 

fresh waters that are N-limited or N and phosphorous (P) co-limited (ISA 3.3.3.2). In a metaanalysis that included 653 datasets, Elser et al. (2007) found that N-limitation occurred as 

frequently as P-limitation in freshwater ecosystems. Additional factors that govern the sensitivity 

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of ecosystems to nutrient enrichment from N deposition include rates and form of N deposition, 

elevation, climate, species composition, plant growth rate, length of growing season, and soil N 

retention capacity. (ISA 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 N deposition compared to acidification. 

2.2.3 What is the magnitude of ecosystem responses to nitrogen deposition? 

2.2.3.1 Terrestrial 

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 N 

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

sensitive to perturbation caused by N additions (LeBauer and Treseder, 2008). Effects are most 

likely to occur where areas of relatively high atmospheric N deposition intersect with N-limited 

plant communities. The alpine ecosystems of the Colorado Front Range, chaparral watersheds of 

the Sierra Nevada, lichen and vascular plant communities in the San Bernardino Mountains and 

the Pacific Northwest, and the southern California coastal sage scrub (CSS) community are 

among the most sensitive terrestrial ecosystems. There is growing evidence that existing 

grassland ecosystems in the western United States are being altered by elevated levels of N 

inputs, including inputs from atmospheric deposition (Clark and Tilman, 2008; Suding et al., 

2005). 

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

assessed in terms of the degree of NO3

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 leaching from watershed soils into ground water or 

surface water. Stoddard (1994) estimated the number of surface waters at different stages of 

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

60% were in Stage 1 or Stage 2 of N 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 

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hardwood forests, alpine meadows, arid and semi-arid lands, and grassland ecosystems (ISA 

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The REA used published research results (REA 5.3.1 and ISA Table 4.4) to identify 

meaningful ecological benchmarks associated with different levels of atmospheric nitrogen 

deposition. These are given by Figure 2-8. 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 NADP 

monitoring data. This map, reproduced in Figure 2-9, depicts the sites where empirical effects of 

terrestrial nutrient enrichment have been observed and site proximity to elevated atmospheric N 

deposition. 

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Figure 2-8. Benchmarks of atmospheric nitrogen deposition for several ecosystem 

indicators (REA 5.3.1.2) MCF-Mixed Conifer Forest, CSS-Coastal Sage 

Scrub

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1 

2 3 4 5 6 7 

Figure 2-9 Observed effects from ambient and experimental atmospheric nitrogen deposition loads in r

using CMAQ 2002 modeling results and NADP monitoring data. Citations for effect results are from 

4.4 (U.S. EPA, 2008) 1= Fenn et. al. (2008), 2=Weiss (1999), 3=Bytnerowicz and Fenn (1996), 4=Fenn et al. (2000), 

Fenn (2004), 6=Jones et al. 2004, 7=Baron (2006), 8=Baron et al. (2000), 9=Gotelli and Ellison (2002), 10=Stoddard 

11=Egerton Warburton and Allen (2000), 12=Brooks (2003), 13=Baez et al. (2007), 14=Bowman et al. (2006), 15=B

(1995), 16=Rueth et al. (2003), 17=DeWalle et al. (2006), 18=Clark and Tillman (2008), 19=Rueth et al. 2003

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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 

CCS community and Mixed Conifer Forest (MCF) (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’s nutrient-sensitive acidophyte lichen communities, fine-root biomass in Ponderosa 

pine, and leached nitrate in receiving waters. 

The ecological benchmarks that were identified for the CSS and the MCF are 

included in the suite of benchmarks identified in the ISA (ISA 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: 

 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 

 10 kg N/ha/yr – mycorrhizal community changes 

For the MCF community, the following ecological benchmarks were identified: 

 3.1 kg N/ha/yr – shift from sensitive to tolerant lichen species 

 5.2 kg N/ha/yr – dominance of the tolerant lichen species 

 10.2 kg N/ha/yr – loss of sensitive lichen species 

 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. 

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% 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. 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. 2002 CMAQ/NADP nitrogen deposition data exceeds the 3.1 kg 

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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. 

2.2.3.2 Freshwater 

 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 N deposition. As noted previously, Elser et al. (2008) found that Nlimitation occurs as frequently as P-limitation in freshwater ecosystems (ISA 3.3.3.2). 

Recently, a comprehensive study of available data from the northern hemisphere surveys 

of lakes along gradients of N deposition show increased inorganic N concentration and 

productivity to be correlated with atmospheric N deposition (Bergström and Jansson 

2006). The results are unequivocal evidence of N limitation in lakes with low ambient 

inputs of N, and increased N concentrations in lakes receiving N solely from atmospheric 

N deposition (Bergström and Jansson, 2006). These authors suggested that most lakes in 

the northern hemisphere may have originally been N-limited, and that atmospheric N 

deposition has changed the balance of N and P in lakes. 

Available data suggest that the increases in total N 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 N deposition that was only about 1.5 kg 

N/ha (Baron, 2006). Similar changes inferred from lake sediment cores of the Beartooth 

Mountains of Wyoming also occurred at about 1.5 kg N/ha deposition (Saros et al., 

2003). Pre-industrial inorganic N deposition is estimated to have been only 0.1 to 0.7 kg 

N/ha based on measurements from remote parts of the world (Galloway et al., 1995; 

Holland et al., 1999). In the western U.S., pre-industrial, or background, inorganic N 

deposition was estimated by (Holland et al., 1999) to range from 0.4 to 0.7 kg N/ha/yr. 

Eutrophication effects from N 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 N deposition effects is expected 

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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 (Fenn et al., 1996), the Front Range of Colorado 

(Baron et al., 1994; Williams et al., 1996), the Allegheny mountains of West Virginia 

(Gilliam et al., 1996), the Catskill Mountains of New York (Murdoch and Stoddard, 

1992; Stoddard, 1994), the Adirondack Mountains of New York (Wigington et al., 1996), 

and the Great Smoky Mountains in Tennessee (Cook et al., 1994) (ISA 3.3.8). 

2.2.3.3 Estuaries 

In contrast to terrestrial and freshwater systems, atmospheric N load to estuaries 

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

of N greatly exceed N deposition. In estuaries, N-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 United States. 

What is the nature of estuary responses to reactive nitrogen deposition? 

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

deposition and the biogeochemical cycling of N and carbon (C) in estuaries (ISA 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 (REA 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% (REA 5.1.1). 

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

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

ecosystems (ISA 4.3.4.2 and 3.3.5.4). Atmospheric and non-atmospheric sources of N 

contribute to increased phytoplankton and algal productivity, leading to eutrophication. 

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Shifts in community composition, reduced hypolimnetic DO, decreases in biodiversity, 

and mortality of submerged aquatic vegetation are associated with increased N deposition 

in estuarine systems. 

 What types of ecosystems are sensitive to such effects? How are these 

responses affected by atmospheric, ecological, and landscape factors? 

Because the productivity of estuarine and near shore marine ecosystems is 

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

of N deposition (ISA 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 (Bricker et al., 2007). 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) (Bricker et al., 2007) was used to 

evaluate eutrophication due to atmospheric loading of nitrogen. 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) 

(Bricker et al., 2007). 

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 (REA 5.2.4.1 and 5.2.4.2). 

These results were then considered with SPAtially Referenced Regression (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% of the total atmospheric nitrogen 

deposition be eliminated, but considerably more nitrogen from other sources as well must 

be controlled (REA section 5.2.7.2). In the Potomac River estuary, a 78% decrease of 

total nitrogen could move the EI score from Bad to Poor (REA 5.2.7.1). The results of 

this analysis indicated decreases in atmospheric deposition alone could not eliminate 

coastal eutrophication problems due to multiple non-atmospheric nitrogen inputs (REA 

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7.3.3). However, by decreasing atmospheric contributions, it may help avoid the need for 

more costly controls on nitrogen from other sources. In addition, the somewhat arbitrary 

discreteness of the EI scale can mask the benefits of decreases in nitrogen between 

categories. 

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In general, estuaries tend to be N-limited (Elser et al., 2008), and many currently 

receive high levels of N input from human activities to cause eutrophication (Howarth et 

al., 1996; Vitousek and Howarth, 1991). Atmospheric N loads to estuaries in the U.S. are 

estimated to range from 2-8% for Guadalupe Bay, TX on the lowest end to as high as 

72% for St Catherines-Sapelo estuary, GA (Castro et al., 2003). The Chesapeake Bay is 

an example of a large, well-studied and severely eutrophic estuary that is calculated to 

receive as much as 30% of its total N load from the atmosphere. 

What is the magnitude of ecosystem responses to eutrophication? 

There is a scientific consensus that nitrogen-driven eutrophication in shallow 

estuaries has increased over the past several decades and that the environmental 

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

widespread occurrence (Paerl et al., 2001). For example, the frequency of phytoplankton 

blooms and the extent and severity of hypoxia have increased in the Chesapeake Bay 

(Officer et al., 1984) and Pamlico estuaries in North Carolina (Paerl et al., 1998) and 

along the continental shelf adjacent to the Mississippi and Atchafalaya rivers’ discharges 

to the Gulf of Mexico (Eadie et al., 1994). 

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

of the assessed systems had moderate to high overall eutrophic conditions and generally 

received the greatest N loads from all sources, including atmospheric and land-based 

sources (Bricker et al., 2007). Most eutrophic estuaries occurred in the mid-Atlantic 

region and the estuaries with the lowest degree of eutrophication were in the North 

Atlantic (Bricker et al., 2007). Other regions had mixtures of low, moderate, and high 

degrees of eutrophication (ISA 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 (ISA 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, 

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and traps sediments (ISA 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. 

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2.2.4 What are the 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 and summarized in section 2.1.4 of this chapter (REA 

7.0). 

2.2.4.1 Aquatic 

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 RETrival (STORET) database, NOAA’s Estuarine Drainage Areas data, 

and EPA’s water quality standards nutrient criteria for rivers and lakes (REA Appendix 

6, 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 

(REA 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 (REA 7.3.3). 

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Ecological thresholds are identified for CSS and MCF and these data are 

considered to be of high quality, however, the ability to extrapolate these data to larger 

regional areas is limited (REA 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. 

2.3 WHAT ECOLOGICAL EFFECTS ARE ASSOCIATED WITH GASPHASE NOX AND SOX? 

Acidifying deposition and nitrogen enrichment are the main focus of this policy 

assessment; however, there are other known ecological effects are attributed to gas-phase 

NOx and SOx. Acute and chronic exposures to gaseous pollutants such as sulfur dioxide 

(SO2), nitrogen dioxide (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 

NOx or SOx, 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 N and S. Most phototoxic effects associated with gas 

phase NOx and SOx occur at levels well above ambient concentrations observed in the 

U.S. (ISA 3.4.2.4). 

2.3.1 What is the nature of ecosystem responses to gas-phase nitrogen and sulfur? 

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

effects (ISA 4.4). The evidence is sufficient to infer a causal relationship between 

exposure to SO2 and injury to vegetation (ISA 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 (ISA 4.4.2 and 3.4.2.2). At sufficient 

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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 (ISA 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. 

2.3.2 What types of ecosystems are sensitive to such effects? How are these 

responses affected by atmospheric, ecological, and landscape factors? 

Vegetation in ecosystems near sources of gaseous NOx and SOx or where 

ambient concentrations of SO2, NO, NO2, PAN and HNO3 are higher 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 (ISA 3.4.2). The functional relationship between ambient concentrations of gas 

phase NOx and SOx and specific plant response are impacted by internal factors such as 

rate of stomatal conductance and plant detoxtification mechanisms, and external factors 

including plant water status, light, temperature, humidity, and pollutant exposure regime 

(ISA 3.4.2). 

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

gas phase as well as to stomatal aperature. The aperature 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 (ISA 3.4.2). 

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

conditions affecting stomatal conductance in vascular plants (REA 6.4.1). For example, 

damage to lichens from SO2 exposure include decreases in photosynthesis and 

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

inhibition of nitrogen fixation, decreased K+

 absorption, and structural changes (Belnap 

et al., 1993; Farmer et al., 1992, Hutchinson et al., 1996). 

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2.3.3 What is the magnitude of ecosystem responses to gas phase effects of NOx 

and SOx? 

The phytotoxic effects of gas phase NOx and SOx are dependent on the exposure 

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

vegetation associated with NOx and SOx, 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, 

with fruticose lichens being more susceptible to SO2 than many foliose and crustose 

species (Hutchinson et al., 1996). 

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

NO2, PAN and HNO3 at concentrations currently observed in the United States with the 

exception of some lichen species (ISA 4.4). Past and current HNO3 concentrations may 

be contributing to the decline in lichen species in the Los Angeles basin (Boonpragob and 

Nash 1991; Nash and Sigal, 1999; Riddell et al., 2008). PAN is a very small component 

of nitrogen deposition in most areas of the United States (REA 6.4.2). Current 

deposition of HNO3 is contributing to N saturation of some ecosystems close to sources 

of photochemical smog (Fenn et al., 1998) such as the MCF’s of the Los Angeles basin 

mountain (Bytnerowicz et al., 1999). Most phototoxic effects associated with gas phase 

NOx and SOx occur at levels well above ambient concentrations observed in the U.S. 

(ISA 3.4.2.4). 

23 2.4 SUMMARY 

In summary, NOx and SOx in the atmosphere contribute to effects on individual 

species and ecosystems through direct contact with vegetation, and more significantly 

through deposition to sensitive ecosystems. The ISA concludes that the evidence is 

sufficient to conclude causal relationships between acidifying deposition of N and S and 

effects on freshwater aquatic ecosystems and terrestrial ecosystems, and between 

nitrogen nutrient enrichment and effects on sensitive terrestrial and freshwater aquatic 

ecosystems. The ISA also concludes that a causal relationship is supported between 

nitrogen nutrient enrichment and effects on estuarine ecosystems; however, the 

September, 2010 2-35 Do Not Quote or Cite 
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The REA provides additional support that under recent conditions, deposition 

levels have exceeded benchmarks for ecological indicators of acidification and nutrient 

enrichment that indicate that effects are likely to be widespread in lakes and streams 

within sensitive ecosystems. 

September, 2010 2-36 Do Not Quote or Cite 
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3.1 How do we characterize adversity to public welfare? What are the relevant factors 

and how are they addressed in this document? 

Characterizing a known or anticipated adverse effect to public welfare is an important 

component of developing any secondary NAAQS. According to the Clean Air Act, 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 

adversity to public welfare as deriving from 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, PM (relating 

to visibility), as well as in initial reviews of effects from lead deposition. Because NOy and SOx 

are deposited from ambient sources into ecosystems where they affect changes to organisms, 

populations and ecosystems, the concept of adversity to public welfare as related to impacts on 

the public from alterations in structure and function of ecosystems is appropriate for this review. 

Other information that may be helpful to consider includes the role of critical loads and 

ecosystem service impacts as benchmarks or measures of impacts on ecosystems that may affect 

public welfare. Ecosystem services can be related directly to concepts of public welfare to 

inform discussions of societal adverse impacts. In subsequent sections we will discuss each of 

these concepts as they relate to adversity. 

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3.1.1.1 Ozone NAAQS Review 

The evaluation of adversity from a public welfare perspective in the context of ozone and 

particulate matter (PM) are relevant to this current review. Both ozone and PM have 

documented effects on ecological receptors. These criteria pollutants are being reviewed on a 

schedule as part of the NAAQS process. The ozone secondary standard is currently under 

reconsideration from the 2008 ruling with a proposal was published January 6, 2010. The final 

Policy Assessment for PM is being developed and is expected to be finalized in the fall of 2010. 

For the purposes of the reconsideration of the secondary standard for ozone welfare 

effects of ozone are primarily limited to vegetation. These effects begin at the level of the 

individual cell and accumulate up to the level of whole leaves and plants. If effects occur on 

enough individual plants within the population, communities and ecosystems may be impacted. 

Prior to the 2008 ozone review, Ozone vegetation effects were classified as either “injury” or 

“damage” (FR 72 37889). “Injury” was defined as; encompassing all plant reactions, including 

reversible changes or changes in plant metabolism, quality or reduced growth that does not 

impair the intended use of the plant while “damage” includes those injury effects that reach 

sufficient magnitude as to decrease or impair the intended use of the plant (FR 72 37890). The 

“intended use” of the plant was imbedded with the concept of adversity to public welfare. 

Ozone-associated “damage” was considered adverse if the intended use of the plant was 

compromised (i.e. crops, ornamentals, plants located in Class I areas). Effects of ozone on 

single plants or species grown in monocultures such as agricultural crops and managed forests 

were evaluated without consideration of potential effects on natural forests or entire ecosystems. 

In the 2008 rulemaking, EPA expanded the characterization of adversity beyond the 

individual plant level and this language is continued in the 2010 ozone reconsideration. The 

2008 final rule and 2010 proposal conclude that a determination of what constitutes an “adverse” 

welfare effect in the context of secondary NAAQS review can appropriately occur by 

considering effects at higher ecological levels (populations, communities, ecosystems) as 

supported by recent literature. The ozone review uses the example of the construct presented in 

Hogsett et al. (1997) as a model for assessing risks to forests. This study suggests that adverse 

effects could be classified into one or more of the following categories: (1) economic production, 

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(2) ecological structure, (3) genetic resources, and (4) cultural values”. Another recent 

publication, “A Framework for Assessing and Reporting on Ecological Condition: an SAB 

report” (Young and Sanzone, 2002) provides additional support for expanding the consideration 

of adversity beyond the species level and at higher levels by making explicit the linkages 

between stress-related effects at the species level and at higher levels within an ecosystem 

hierarchy (See Figure 3-1). 

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Figure 3-1. Common anthropogenic stressors and the essential ecological attributes they 

affect. Modified from Young and Sanzone (2002) 

In the 2008 ozone NAAQS review and current ozone NAAQS proposal, the 

interpretation of what constitutes an adverse effect on public welfare can vary depending on the 

location and intended use of the plant. The degree to which O3-related effects are considered 

adverse to public welfare depends on the intended use of the vegetation and its significance to 

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public welfare (73 FR 16496). Therefore, effects on vegetation (e.g., biomass loss, foliar injury, 

impairment of intended use) may be judged to have a different degree of impact on public 

welfare depending, for example, on whether that effect occurs in a Class I area, a city park, 

commercial cropland or private land. 

In the proposed ozone reconsideration in 2010 the Administrator has found that the types 

of information most useful in informing the selection of an appropriate range of protective levels 

is appropriately focused on information regarding exposures and responses of sensitive trees and 

other native species that occur in protected areas such as Class I areas or on lands set aside by 

States, Tribes and public interest groups to provide similar benefits to the public welfare. She 

further notes that while direct links between O3 induced visible foliar injury symptoms and other 

adverse effects (e.g., biomass loss) are not always found, visible foliar injury in itself is 

considered by the National Park Service (NPS) to affect adversely air quality related values 

(AQRV) in Class I areas. While the Administrator recognizes that uncertainty remains as to what 

level of annual tree seedling biomass loss when compounded over multiple years should be 

judged adverse to the public welfare, she believes that the potential for such anticipated effects 

should be considered in judging to what degree a standard should be precautionary (73 FR 

16496). The range of proposed levels from 7 – 15 ppb includes at the maximum level of 15 ppb 

protection of approximately 75% of seedlings from more than 10% biomass loss. 

3.1.2 Other EPA Programs and Federal Agencies 

 Various federal laws and policies exist to protect ecosystem health. How other federal 

agencies and EPA offices consider ecosystem effects in carrying out their programs can help 

inform the Administrator when she evaluates the adversity of ecosystem impacts on public 

welfare. From the 1996 National Acid Precipitation Assessment Program Report to Congress: “ 

The 1990 Clean Air Act Amendments require that the National Acid Precipitation Assessment 

Program (NAPAP) prepare biennial reports to Congress, and that “every four years ... the report 

... shall include the reduction in deposition rates that must be achieved in order to prevent 

adverse ecological effects” (Public Law 101-549, Title IX, Section 903 (j)(3)(F)(i), codified as 

amended at 42 USC §7403(j)(3)(F)(I)). Although the term adverse ecological effects is not 

specifically defined in the Clean Air Act Amendments, a working definition can be derived from 

relevant statements at various locations in the statute. Congress expresses its concern with 

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ecological components (the scope is broad and inclusive, since ecology encompasses the 

interrelationships of organisms and their environment) in the preceding subsection (E) of the 

statute. That subsection requires reporting on “the status of ecosystems (including forest and 

surface waters) ... affected by acid deposition ... including changes in surface water quality and 

forest and soil conditions ... [and] high elevation watersheds” (42 USC §7403(j)(3)(E)(iii-v)). 

The adverse effects of concern to Congress, as evidenced in its findings and declaration of 

purpose, are the “dangers to the public health and welfare ... including injury ... damage ... and 

... deterioration” (42 USC §7401(a)). Based on the intent of Congress, as expressed above and 

elsewhere in the Clean Air Act, and shaped by indications of intent expressed in other relevant 

environmental statutes and regulations, the following working definition of adverse ecological 

effects has been derived and is used in the preparation of the NAPAP report: 

any injury (i.e., loss of chemical or physical quality or viability) to any 

ecological or ecosystem component, up to and including at the regional 

level, over both long and short terms. Similarly, adverse effects for other 

areas of concern addressed in this report—i.e., visibility, materials, and 

human health—consist of loss of quality up to and including at the 

regional level, over both long and short terms.” 

As another example, an effect may be considered adverse to public welfare if it 

contributes to the inability of areas to meet water quality objectives as defined by the Clean 

Water Act. The following federal statutes and policies may prove helpful to consider. 

 3.1.21. Prevention of Significant Deterioration Program 

The Clean Air Act’s Prevention of Significant Deterioration (PSD) program (42 U.S.C. 

7470) purposes include to “preserve, protect and enhance the air quality in national parks, 

wilderness areas and other areas of natural, recreational, scenic or historic value . . . .” Also, the 

PSD program charges the Federal Land Managers, including the NPS, with “. . . an affirmative 

responsibility to protect the air quality related values . . . “within federal Class I lands. (42 U.S.C. 

7475(d)(2)(B)). 

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3.1.2.2 EPA Office of Water

Section 101 of the Clean Water Act (CWA) (Declaration of Goals and Policy) states that 

the objective of the CWA is to restore and maintain the chemical, physical, and biological 

integrity of the Nation’s waters and to attain, where possible, water quality that protects fish, 

shellfish, wildlife and provides for water-based recreation. 

The CWA also authorizes EPA to develop water quality criteria as a guide for the states 

to set water quality standards to protect aquatic life. In consideration of acidification effects, 

EPA’s Redbook, Quality Criteria for Water, published originally in 1976, recommends that 

alkalinity be 20 mg/l\L or more as CaCO3 for freshwater aquatic life except where natural 

concentrations are less. Alkalinity is the sum total of components in the water that tend to 

elevate the pH of the water above a value of about 4.5. 

As mentioned in the Redbook, alkalinity is expressed as CaCO3 in mg/L. Alkalinity 

differs slightly from ANC in that ANC includes other buffering compounds (Na, Mg, and K) as 

well and includes buffering capacity of particulates in the water sample. Since alkalinity is 

expressed as mg/L and ANC is expressed as μeq/L, alkalinity must be multiplied by 20 to be 

converted to μeq/L. Thus a recommended criterion of 20 mg/L alkalinity is equivalent to an 

ANC of 400 μeq/L. 

3.1.2.3 National Park Service

The National Park Service (NPS) is responsible for the protection of all resources within 

the national park system. These resources include those that are related to and/or dependent 

upon good air quality, such as whole ecosystems and ecosystem components. The NPS, in its 

Organic Act (16 U.S.C. 1), is directed to conserve the scenery, natural and historic objects and 

wildlife and to provide for the enjoyment of these resources unimpaired for current and future 

generations. 

The Wilderness Act of 1964 asserts wilderness areas will be administered in such a 

manner as to leave them unimpaired and preserve them for the enjoyment of future generations. 

NPS Management Policies (2006) guide all NPS actions including natural resources 

management. In general, the NPS Management Policies reiterate the NPS Organic Act’s 

mandate to manage the resources “unimpaired.” 

September, 2010 3-6 Do Not Quote or Cite 
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On endangered species, Title 16 USC Chapter 35 Section 1531 states “The Congress 

finds and declares that— these species of fish, wildlife , and plants are of esthetic , ecological, 

educational, historical, recreational, and scientific value to the Nation and its people and that all 

Federal departments and agencies will use their authorities to conserve threatened and 

endangered species. 

The United States Fish and Wildlife Service (FWS) manages the National Wildlife 

Refuge System lands to “...ensure that the biological integrity, diversity, and environmental 

health of the Systems are maintained for the benefit of present and future generations of 

Americans.” 16 U.S.C. Section 668dd(a)(4)(B)(1997). 

3.1.2.4 U.S. Forest Service

The National Forest units are managed consistent with Land and Resource Management 

Plans (LRMPs) under the provisions of the National Forest Management Act (NFMA). 16 

§U.S.C. 1604 (1997). LRMPs are, in part, specifically based on recognition that the National 

Forests are ecosystems and their management for goods and services requires an awareness and 

consideration of the interrelationships among plants, animals, soil, water, air, and other 

environmental factors within such ecosystems. 36 C.F.R. §219.1(b)(3) 

Any measures addressing Air Quality Related Values (AQRV) on National Forest 

System lands will be implemented through, and be consistent with, the provisions of an 

applicable LRMP or its revision (16 U.S.C. §1604(i)). Additionally, the Secretary of Agriculture 

must prepare a Renewable Resource Program that recognizes the need to protect and, if 

necessary, improve the quality of air resources. 16 U.S.C. §1602(5)(C). 

AQRVs in Wilderness areas may receive further protection by the previously mentioned 

1964 Wilderness Act. For Wilderness Areas in the National Forest System, the Act's 

implementing regulations are found at 36 C.F.R. §293 requiring these Wilderness Areas be 

administered to preserve and protect [their] wilderness character. 

3.1.25Chesapeake Bay Total Maximum Daily Loads 

Under section 303(d) of the Clean Water Act, states, territories, and authorized tribes are 

required to develop lists of impaired waters. These are waters that are too polluted or otherwise 

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degraded to meet the water quality standards set by states, territories, or authorized tribes. The 

law requires that these jurisdictions establish priority rankings for waters on the lists and develop 

TMDLs for these waters. A Total Maximum Daily Load, or TMDL, is a calculation of the 

maximum amount of a pollutant that a waterbody can receive and still safely meet water quality 

standards. EPA is developing a TMDL for the Chesapeake Bay and its tributaries. The 

Chesapeake Bay Program has modeled the level of nitrogen that can reach the Bay and still meet 

the Bay’s water quality standards. The TMDL, with full public participation, will set waste load 

allocations for point source discharges and load allocations for nonpoint sources of nitrogen. Air 

deposition to the Bay and its watershed, as a source category, will have a specific allocation. 

The allocation can be used to calculate the level of ambient air concentrations of reactive 

nitrogen that are likely to meet the deposition allocation. To find the NOy portion of the 

allocation one would subtract the reduced forms from the total allocation. If the total load to the 

Bay of nitrogen from all the allocated source categories remains below the allocations, then the 

Bay is expected to meet the water quality standards, which are set to protect the designated uses 

of the Bay. Since the designated uses are set by the states with public input, not meeting the 

designated uses can be seen as having an adverse effect to public welfare. 

3.1.2.6 Critical Loads 

The term critical load is used to describe the threshold of air pollution deposition that 

causes a specified level of harm to sensitive resources in an ecosystem. A critical load is 

technically defined as “the quantitative estimate of an exposure to one or more pollutants below 

which significant harmful effects on specified sensitive elements of the environment are not 

expected to occur according to present knowledge” (Nilsson and Grennfelt, 1988). The 

determination of when a harmful effect becomes “significant” may be in the view of a researcher 

or through a policy development process. Researchers often use the term “critical loads” to 

describe when particular detrimental effects are realized, as is the case in Figure 2-1. 

Harmful effects due to acidification have been defined here as those that occur below a 

given ANC for aquatic systems and below a given Bc:Al ratio for terrestrial systems. However, 

the level at which an effect becomes adverse to public welfare is determined by the 

Administrator, informed by available scientific information. 

September, 2010 3-8 Do Not Quote or Cite 
1 3.1.2.7 United Nations Economic Commission for Europe (UNECE)

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In many European countries a critical loads framework is used to determine a level of 

damages to ecosystem services from pollution that is legally allowed. These critical loads are 

determined through a policy process. Indeed critical loads have been modeled by individual 

countries and submitted to the UNECE (in cases where countries have not submitted their own 

critical loads those loads have been calculated for them) and are being used to support 

international emissions reduction agreements including the 1999 Gothenburg protocol and the 

National Emission Ceiling Directive of the European Commission. Figure 3-2 shows critical 

loads for eutrophication and acidification that protect 95% of forests, seminatural vegetation or 

surface waters in Europe. For comparison to the U.S maps presented in this document the units 

of deposition convert to a range of 3.2 kg ha-1

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Figure 3-2 European maps of eutrophication (left) and acidification (right) which 

protect 95% of natural areas in 50x50 km2

 European Monitoring and 

Evaluation Programme grid. Red shaded areas illustrate grid cells where 

deposition needs to be lower than 200 eq ha-1

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Source: Critical Load, Dynamic Modelling and Impact Assessment in Europe CCE Status Report 2008 available at 

http://www.pbl.nl/en/publications/2009/Critical-load-dynamic-modelling-and-impact-assessment-inEurope-CCE-Status-Report-2008.html 

The Coordination Center for Effects, a working center for the Working Group on Effects 

of the Convention on Long Range Transboundary Pollution, in the 2008 status report shows 

calculated critical loads based on an ANC target of 20 μeq/L for the protection of brown trout. 

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Individual countries have also set ANC targets for other species for example Norway uses a 

critical limit of 30μeq/L for Atlantic salmon (Jenkins et al, 2003). 

3.2 What are ecosystem services and how does this concept relate to public welfare? 

An additional concept that may be useful in considering the issue of adversity to public 

welfare is ecosystem services. In the next section the concept of ecosystem services, its 

relationship to adversity and public welfare within the context of this review are explained. 

Ecosystem services can be generally defined 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), and supporting (nutrient 

cycling) (MEA 2005). Conceptually, changes in ecosystem services may be used to aid in 

characterizing a known or anticipated adverse effect to public welfare. In the context of this 

review, ecosystem services may also aid in assessing the magnitude and significance to the 

public of a resource and in assessing how NOy and SOx concentrations and deposition may 

impact that resource. The relationship between ecosystem services and public welfare effects is 

illustrated in Table 3-1. 

September, 2010 3-10 Do Not Quote or Cite 
1 Table 3-1. Crosswalk between Ecosystem Services and Public Welfare Effects 

Public Welfare Effect Ecosystem Service Service Category 

Soils Nutrient Cycling Supporting 

Water Drinking water, Recreation, 

Aesthetic 

Provisioning, Cultural 

Crops Food, Fuel Production, Forest 

Products 

Provisioning 

Vegetation Food, Recreation, Aesthetic, 

Nonuse 

Provisioning, Cultural 

Wildlife Recreation, Food, Nonuse Cultural, Provisioning 

Climate Climate Control Regulating 

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Wellbeing 

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*All ecosystem services contribute to personal comfort and wellbeing. 

EPA has 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” (US EPA 2006). Additionally Executive Order 12866 requires 

a regulatory Impact Analysis for any rule considered “economically significant” and defines 

significant as a rule having $100 million or more in impacts. Though this is not a definition 

specifically for use in the NAAQS process it may be a useful one in considering the scope of 

ecosystem services and the effects of air pollutants upon those services. Especially important is 

the acknowledgement that it is difficult to measure and/or monetize the goods and services 

supplied by ecosystems. Valuing ecological benefits, or the contributions to social welfare 

derived from ecosystems, can be challenging as noted in EPA’s Ecological Benefits Assessment 

Strategic Plan (US EPA 2006) and the Science Advisory Board report “Valuing the Protection of 

Ecological Systems and Services” (US EPA, 2009). It can be informative in characterizing 

adversity to public welfare to attempt to place an economic valuation on the set of goods and 

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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. The stepwise concept leading to the 

valuation of ecosystem services is graphically depicted in Figure 3-3. 

Figure 3-3. Representation of the benefits assessment process indicating where some 

ecological benefits may remain unrecognized, unquantified, or unmonetized. 

(Source: EBASP USEPA 2006). 

A conceptual model integrating the role of ecosystem services in characterizing known or 

anticipated adverse effects to public welfare is shown in Figure 3-3. Under Section 109 of the 

CAA, the secondary standard is to specify a level of air quality that is requisite to protect public 

welfare. For this review, the relevant air quality indicator is interpreted as ambient NOy and SOx

concentrations that can be linked to levels of deposition for which there are ecological effects 

that are adverse to public welfare. The case study analyses (described in Chapters 4 and 5 of the 

REA and summarized in Chapter 2 of this document) link deposition in sensitive ecosystems 

(e.g., the exposure pathway) to changes in a given ecological indicator (e.g., for aquatic 

acidification, changes in acid neutralizing capacity [ANC]) and then to changes in ecosystems 

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and the services they provide (e.g., fish species richness and its influence on recreational 

fishing). To the extent possible for each targeted effect area, ambient concentrations of nitrogen 

and sulfur (i.e., ambient air quality indicators) were linked to deposition in sensitive ecosystems 

(i.e., exposure pathways), and then deposition was linked 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 their public welfare effects (e.g., recreational fishing). 

We recognize that there is a certain amount of natural change in ecosystems over time that can 

affect the level of acidity and the response of the ecosystem to additional acid and nutrient 

inputs. However, this review is focused on the impact of anthropogenic nitrogen and sulfur 

deposition given the existing state of non-anthropogenically determined ecosystem 

characteristics, and as such we essentially hold these other factors as “fixed” for the purposes of 

the review. 

Knowledge about the relationships linking ambient concentrations and ecosystem 

services can be used to inform a policy judgment on a known or anticipated adverse public 

welfare effect. The conceptual model outlined for aquatic acidification in Figure 3-3 can be 

modified for any targeted effect area where sufficient data and models are available. 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, 

recreational viewing and aesthetics—would be classified as public welfare effects. This 

information can then be used by the Administrator to determine whether or not the changes 

described are adverse to public welfare. In subsequent sections these concepts are applied to 

characterize the ecosystem services potentially affected by nitrogen and/or sulfur for each of the 

effect areas assessed in the REA. 

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Figure 3-4. Conceptual model showing the relationships among ambient air quality 

indicators and exposure pathways and the resulting impacts on ecosystems, 

ecological responses, effects and benefits to characterize known or 

anticipated adverse effects to public welfare. 

These concepts can also be applied to the programs described in section 3.1. National 

parks represent areas of nationally recognized ecological and public welfare significance, which 

are afforded a higher level of protection. Therefore, staff has also focused on air quality and 

deposition in the subset of national park sites and important natural areas. The spatial 

relationships between sensitive regions, Class 1 areas and nitrogen deposition levels are 

illustrated in Figures 3-5 and 3-6. Please note that the scale of deposition levels is different for 

the two maps to allow greater differentiation of the deposition in the western U.S. 

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2 Figure 3-5. Locations of Eastern U.S. Public Lands relative to deposition of nitrogen and 

3 sulfur in sensitive aquatic areas. Source 2005 CMAQ and NADP.

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3 Figure 3-6. Location of Western U.S. Public Lands relative to deposition of nitrogen and 

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1 3.3 Applying Economic Valuation to Ecosystem Services 

2 As discussed earlier in this document, a secondary NAAQS is required to be set at the 

3 “level(s) of air quality necessary to protect the public welfare from any known or anticipated 

4 adverse effects”. As part of the effort to determine the standard, EPA linked the changes in the 

5 ambient air concentrations of NOy and SOx to the changes in ecosystem services and ultimately 

6 to changes in public welfare (U.S. EPA, 2009). The difficulty in the monetization for ecosystem 

7 services has been previously emphasized. This difficulty necessitates focusing on a subset of 

8 services for economic valuation. And although economics on its own cannot determine what 

9 level of impact on public welfare is “adverse,” economics can be helpful in the context of a 

10 secondary NAAQS for determining the degree to which improvements are beneficial to public 

welfare and illustrating and aggregating those impacts.1 11 

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13 3.3.1 Ecosystem Services and Links to Public Welfare 

14 An ecosystem service framework provides a structure to measure changes in public 

15 welfare from changes in ecosystem functions affected by air pollution. EPA’s Risk Assessment 

16 for this rulemaking defines ecosystem services as “the ecological processes or functions having 

17 monetary or nonmonetary value to individuals or society at large” (EPA 2009) The discipline of 

18 economics provides a useful approach for summarizing how the public values changes in the 

19 services provided by the environment. An ecosystem services framework (with or without 

20 valuation) can provide measures of changes in public welfare. 

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3.3.2 Economics as a Framework to Illustrate Changes in Public Welfare 

Economics can provide a framework to illustrate how public welfare changes in response 

to changes in environmental quality by quantitatively linking changes in ecosystem services to 

preferences. Economics assumes that the choices that individuals make reflect their preferences 

over certain outcomes and that, generally speaking, they will make choices that, in expectation, 

will make them as well off as possible given their resources. In economics revealed and stated 

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 Section 109 of the Clean Air Act forbids consideration of the compliance costs of reducing pollution when setting 

a NAAQS. However, there is no prohibition regarding the consideration of the monetized impacts of welfare effects 

occurring due to levels of pollution above alternative standards in evaluating the adversity of the impacts to public 

welfare. Ecosystem services can be characterized as a method of monetizing the impacts of the air pollution. 

Although a separate regulatory document quantifying the costs and benefits of attaining a NAAQS is prepared 

simultaneously, this document is not considered when selecting a standard. 

September, 2010 3-17 Do Not Quote or Cite 
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preference methods are used to observe the choices individuals make to understand the outcomes 

individuals prefer. What individuals are willing to give up for an outcome is their willingness-topay (WTP) for that outcome. An example of an outcome is an improvement in an ecosystem 

service. Often, to provide comparability to other goods and services, in economics these 

tradeoffs are framed relative to dollars for convenience. 

Economics could inform the Administrator by valuing and characterizing the changes in 

public welfare from changes in the quantity and quality of ecosystem services. Overall, this 

assessment intends to characterize changes in ecosystem services from a scientific perspective 

using effects on ecosystem structures and functions or ecosystem integrity. Economics then 

estimates the effect on public welfare of these changes in the quantity and quality of ecosystem 

services using willingness to pay as a measure of this effect. For example, a decrease in a 

particular bird species can be characterized by its effect on the ecosystem’s structure and 

function, while from an economic perspective, the effects would be based on the impact on 

public welfare or the value the public places on that species. A simple example is a comparison 

between a decrease in a bird species that is relatively unknown compared to a decrease in a very 

prominent species (e.g. bald eagle). The public is likely to have a higher WTP to avoid the latter, 

and thus the decrease would affect the public welfare more, even if the changes in the two bird 

species generally have the same impact on an ecosystem’s structure or function. 

19 There are important complications with using preferences to understand the effect of 

20 pollution on public welfare. For example, while the field of economics generally assumes that 

21 public preferences are the paramount consideration; care must be taken that these preferences 

22 may change when the public receives new information. Evaluation of public preferences should 

23 take place under conditions of full information. If individuals do not understand how pollution 

24 will affect ecosystem services, or even how those ecosystem services affect their quality of life, 

25 then they will have a difficult time valuing changes in those services. Similarly, it may be very 

26 difficult and time-consuming for individuals to learn and understand how changes in particular 

27 ecosystem services may affect them, in part because typically there are significant 

28 interdependencies within an ecosystem. Because of this complexity, individuals may implicitly 

29 value a species, or habitat, or ecosystem function because it supports an ecosystem service that 

30 they do clearly value. Furthermore, the public also has limited understanding regarding 

31 irreversibilities, tipping points, and other more complex aspects of ecosystems, which limits the 

September, 2010 3-18 Do Not Quote or Cite 
ability to adequately value these ecosystems.2 1 In addition, where and when a change in an 

2 ecosystem takes places is crucial for characterizing the associated change in an ecosystem 

3 service, and will also affect the value the public places on that change. 

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3.3.3 The Role of Economics in Defining “Adversity” 

If economic valuation can establish a significant loss to public welfare, then this can 

provide strong support for a determination of an “adverse” effect. However, there is neither an 

economic definition of how much loss in public welfare is adverse nor an economic definition of 

adversity. While an economist might consider a particular scenario adverse because it might 

imply some harm or potential for improvement, there is no specific threshold level when a loss in 

welfare (e.g. loss in dollars) becomes adverse. An individual might be willing to give up some 

of their resources to avoid a threat or negative outcome (i.e., willing to pay to avoid a particular 

outcome). According to economic theory, if an individual is willing to give up something to 

avoid the outcome, then imposing the outcome on the individual must make them worse off, at 

which point an economist might describe the outcome as adverse. However, the amount an 

individual is willing to pay to avoid the outcome may or may not rise to the level of harm that the 

Administrator interprets as “adverse” to public welfare. At the same time, an economic 

valuation that shows that there are substantial damages from current levels of acidification or 

nutrient enrichment would provide strong evidence for finding that current impacts are adverse 

to public welfare. In summary, economic analysis (particularly valuation) can provide useful 

information for the Administrator as to the interpretation of the word “adverse” in the context of 

public welfare, but it does not provide a complete set of information needed to make that 

determination. 

3.3.44 Collective Action as an Indicator of Public Preferences 

Typically economics uses information on willingness to pay for improved environmental 

quality that is gathered from observing individuals’ market behavior (revealed preference) or that 

they provide through surveys (i.e., stated preference methods). The analyses in the following 

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it may be costly to do so. It is possible for individuals to value outcomes that are irreversible or result in discrete 

changes (i.e., tipping points) in the quality and quantity of ecosystem services. Avoiding irreversible outcomes 

should be and are more valued by individuals than outcomes that are not irreversible (Arrow and Fischer, 1974). 

September, 2010 3-19 Do Not Quote or Cite 
sections use revealed and stated-preference information to quantify a portion of the social costs 

of current levels of acidification and nutrient enrichment. However, the studies supporting these 

analyses evoke specific contexts and thus the findings may not be generalizable across all of 

those affected by acidification or nutrient enrichment.

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 An alternative source of revealed 

information on the damages caused by acidification can be found in the behavior of groups. 

Often groups collectively make choices to engage in activities that improve the collective 

welfare of the group. For example, a community around an acidified lake might undertake 

activities designed to improve the quality of that lake, including purchasing lime, to use as a tool 

to reduce the acidity of the lake. These collective decisions can be used to gain insights into how 

people value improvements to ecosystem services. However, there are many obstacles to 

collective actions, including problems of organization, free ridership and others (Olson 1965) 

that make it difficult to use the actions of organizations to interpret individuals’ preferences. 

In addition to communities, states may also take actions to increase the quality of their 

impaired lakes. Non-Governmental Organizations (NGOs) or advocacy groups, as well may 

organize support for, and/or directly undertake, activities to improve lake and stream quality on 

behalf of its members/donors. How individual’s preferences are expressed through these 

collective actions is discussed below. For brevity, this discussion will focus on collective efforts 

to reduce acidity of lakes and streams by Communities, Nongovernmental Organizations and 

States. 

3.3.4.1 Communities 

In cases where property rights to a resource are well defined, collective action is more 

likely to take place, as individuals have greater ownership and control over the affected resource. 

Rights to use a lake, as well as, mandatory membership in a lake association is often written right 

into the deed of properties which abut or surround a lake, giving these property owners more 

control over the resource. This mechanism of granting rights and responsibilities to the lake 

encourages better management of the lake resource by remedying unrestricted access and free 

rider problems. This coupling of the costs of resource improvements with their benefits 

encourages individuals to maintain the quality of the resource. 

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 Even in the case where the existing studies provide a reliable characterization of the effects of acidification or 

nutrient enrichment on a limited number of individuals, it is advisable to make use of corroborating data and studies 

when such information is available. 

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There have been several documented instances where communities (particularly 

Homeowners Associations) have spent time and money to improve the quality of a lake. These 

include actions to combat acidity, eutrophication, invasive species (e.g. Zebra Mussels) and other 

problems. The Lake Wononscopomuc Association in Salisbury, Connecticut is a typical 

example (Mayland 2009.) They spend their own funds to hire scientific consultants to survey 

and test the lake water (for e.g. pH., dissolved oxygen, visibility, and many other factors related 

to the lake’s condition) and recommend management strategies to improve the quality of the 

lake. Likewise, in Georgia, the Berkeley Lake Homeowners Association (BLHA) is a non-profit 

homeowner association dedicated to protecting Berkeley Lake. BLHA is typical of many other 

home owners associations with access to a lake, in that they are also concerned with and 

managing acidity, eutrophication, invasive species and a whole host of more mundane upkeep 

and maintenance issues (Hunkapiller 2006.) BLHA recognizes the relationship between lake 

acidity and resident’s enjoyment of the lake’s fishing swimming and aesthetics. 

3.3.4.2 Nongovernmental Organizations

 Nongovernmental Organizations (NGOs) or Advocacy Groups organize individuals and 

smaller groups thereby reducing the transaction costs associated with individual’s desires to 

advance a specific goal. For example, Living Lakes, Inc. (LLI) is a not-for profit organization 

which sponsors an applied aquatic resources restoration demonstration program for acidified 

waters. In the late 1980’s LLI began evaluation of seven different liming technologies on 22 

lakes and 10 streams in 6 states. Lakes and streams have been treated in the states of 

Massachusetts, Rhode Island, New York, Pennsylvania, Maryland and West Virginia (Brocksen 

and Emler 1988.) Likewise, sportsman groups such as Trout Unlimited, as well as, smaller local 

groups, have an interest in improving or maintaining the quality of lakes and streams. Trout 

Unlimited is well known for these activities and is discussed further later on. However, several 

smaller, localized groups also work to decrease aquatic acidification. One of these is the 

Mosquito Creek Sportsman’s Association in Pennsylvania. Mosquito Creek and its main 

tributary Gifford Run were once famous for naturally reproducing wild brook and brown trout. 

However, since the early 1960’s, the pH of the stream steadily declined due to acid rain. As a 

result, wild brook trout, as well as, wild brown trout have substantially declined in the watershed 

(Hoover and Rightnour 2002.) Aerial liming was undertaken as part of an overall watershed 

restoration program that included constructed wetlands, forest liming, and in-stream liming to 

September, 2010 3-21 Do Not Quote or Cite 
improve this fishery and provide increased opportunities for public recreation in the region. 

Fifty tons of lime were applied in the headwaters of Mosquito Creek Watershed. This liming 

project was part of the Mosquito Creek Sportsman’s Association’s efforts to improve the water 

quality of the 90 square mile watershed located in Clearfield and Elk counties. However, while 

the project first phase and the other ongoing phases of the overall restoration project have been 

initiated by the Mosquito Creek Sportsman Association, they received technical support from 

multiple public, private and other non-profit groups.

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the four implementation phases of this project. Costs were based on alkaline deficiencies and the 

additional costs determined for the technologies. Benefits were estimated as returns on direct 

recreational use losses and community willingness-to-pay. Full restoration of the watershed is 

estimated to cost approximately $3.4 million over 15 years, for an annualized cost of $229,000, 

or $5,400 per mile per year for 42 miles of potential improvements. Expected returns range from 

$1.2 million per year for recreational use to $6.1 million per year for total community 

willingness-to-pay. It was concluded that restoration is technically feasible and economically 

beneficial for the Mosquito Creek watershed, and it is recommended that planned projects and 

the remainder of the progressive restoration plan be implemented.” (Hoover and Rightnour 

2002.) 

3.3.4.3 States 

Several states including: Vermont, New Hampshire, New York and Tennessee have 

developed Total Maximum Daily Loads (TMDLs) for lakes impaired for acidification in their 

jurisdictions. As mentioned in the previous section regarding the TMDL for the Chesapeake Bay 

the applicable water quality standards and designated uses are set by the states with public 

participation. Although most states set their standard either by legislation or regulations in at 

least one case, specifically New York, the designated uses and water quality standard are part of 

the state constitution. The Adirondack Forest Preserve is required to be “forever kept as wild 

forest lands”. New York has interpreted this to mean that the waters included in the preserve are 

4 These included: the Pennsylvania State University Environmental Resource Research Institute, Pennsylvania 

Game Commission, Pennsylvania DCNR Bureau of Forestry, Pennsylvania Fish and Boat Commission, 

Pennsylvania Department of Environmental Protection, Pennsylvania Department of Corrections Quehanna Boot 

Camp, Wood Duck Chapter Trout Unlimited, Canaan Valley Institute, Clearfield County Conservation District and 

Pennsylvania USDA Natural Resource Conservation Service.

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required to be kept in natural conditions. To this end New York has chosen to set a tiered TMDL 

that allows interim water quality targets in order to address the reality that some lakes in the 

Adirondacks will naturally have a pH that does not meet the state’s water quality standards. For 

lakes that can meet the standards the state has chosen to set the water quality standard for pH is 

above 6.5. New Hampshire has chosen to set their water quality target at an ANC of 60μeq/L 

that, according to the TMDL document, corresponds to a pH of 6.5. Vermont, in a similar 

process has chosen a target ANC of 50μeq/L. In Tennessee the state faces a similar problem as 

New York in trying to set levels to protect streams Great Smoky Mountains National Park which 

include some naturally acidic streams. Accordingly they have set site specific ANC targets 

where data is available to do so and chosen to target an ANC of 50μeq/L as a default value 

where data is not available. The Tennessee TMDL is a partnership between the state and the 

National Park Service which is sharing the data collection and modeling activities with academic 

institutions. 

In each case the implementation sections of these TMDLs cites the fact that the sources 

of pollution responsible for the degradation of water quality in the named lakes and streams are 

not located within the jurisdiction of the state. Each state has called on EPA to require regional 

or national decreases in acidifying deposition. Vermont goes so far as to say “In short, 

implementation of this TMDL is primarily the responsibility of EPA.... This TMDL sets out 

clear endpoints to guide EPA actions. However, in the absence of vigorous efforts by EPA to 

bring about reductions in acid emissions from out-of-state sources this TMDL will merely have 

been a paper exercise.” 

3.3.4.4 Public-Private Collaborations

In some cases, industry, government and private efforts partner to reduce the acidity of a 

lake or stream. In one such instance in 2005, The U.S. Forest Service used helicopters to apply 

200 tons of limestone sand into the St. Marys River and its tributaries to lower acid levels in one 

of Virginia's prime trout fisheries to mitigate the impacts of acidification until a long-term 

solution to acidification is found. The NGO, Trout Unlimited was one of the partners in the 

liming project, while Dominion Virginia Power provided $10,000 for the liming project 

(Associated Press 2005.) In another partnership, Living Lakes participated in a project in the 

Woods Lake Watershed in the Adirondack region of the state of New York that was coSeptember, 2010 3-23 Do Not Quote or Cite 
sponsored by the Electric Power Research Institute (EPRI.) There are many such examples, 

where all three of these types of groups partner on the same project. 

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3.3.4.5 Using Information on Collective Actions to Estimate Welfare Impacts 

The fact that collective action activities are being undertaken by communities, NGO’s 

and States underscores the fact that there is a societal demand for further improvement to the 

quality of many water bodies which have been impaired by acidic deposition.5

 However, they 

provide insufficient quantitative evidence as to what society willingness to pay to reduce lake 

and stream acidity because it is difficult to separately identify individual preferences from the 

actions of the group. For example, groups suffer from the problem of free-ridership. When there 

are free-riders we know that the activity of the group understates the preferences of the 

community for an improvement in environmental quality, however it is difficult to estimate the 

magnitude of its impact on the total activity of the group. 

3.4 Effects of Acidification and Nutrient Enrichment 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 were linked to an ecological response that was subsequently linked, to the extent 

possible, to associated services. 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 as illustrated in the conceptual model shown in Figure 3-7. 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. 

5 However, one must recognize that often times reducing acidity is often part of a larger effort to generally improve 

the quality of a water body. Therefore, separating out the portion of people’s desire to just to reduce acidity from 

the more general improvements is difficult. 

September, 2010 3-24 Do Not Quote or Cite 
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Figure 3-7. Conceptual model linking ecological indicator (ANC) to affected ecosystem 

services. The red arrows highlight the path to monetization of recreational 

fishing effects. 

While aquatic acidification is the focus of this policy assessment, the other effect areas 

analyzed in the REA still merit some discussion in view of the fact that 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. In next four sections we 

summarize the current levels of specific ecosystem services for aquatic and terrestrial 

acidification, and aquatic and terrestrial nutrient enrichment. We also present results of analyses 

that have attempted to quantify and monetize the harms to public welfare, as represented by 

ecosystem services, due to nitrogen and sulfur deposition. 

For the purposes of the following sections nutrient enrichment refers only to that due to 

NOy deposition. Additionally these sections focus on the detrimental effects of that deposition. 

Staff acknowledges that a certain amount of NOy deposition in managed terrestrial ecosystems 

may have a beneficial effect, specifically increased growth (a fertilization effect). However no 

attempt has been made to quantify those beneficial effects since this document and preceding 

analyses are focused on unmanaged sensitive ecosystems. 

September, 2010 3-25 Do Not Quote or Cite 
1 3.4.1 Evidence for Adversity Related to Aquatic Acidification 

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Acidification primarily affects the ecosystem services that are derived from the fish and 

other aquatic life found in these surface waters (REA, Section 5.2.1.3). 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. For example, although there is evidence that certain 

population subgroups in the northeastern United States, such as the Hmong and Chippewa ethnic 

groups, have particularly high rates of self-caught fish consumption (Hutchison and Kraft, 1994; 

Peterson et al., 1994), it is not known if and how their consumption patterns are affected by the 

reductions in available fish populations caused by surface water acidification.

Inland surface waters support several cultural services, such as recreational fishing, 

aesthetic and educational services; however, Banzhaf et al (2006) has shown that non-use 

services are arguably a significant source of benefits from reduced acidification. The areas of the 

country containing the most sensitive lakes and streams are New England, the Adirondack 

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

region), northern Florida, and the Upper Midwest. Within the Adirondack Mountains 

approximately 8% of the lakes were considered acidic and in the northern Appalachian Plateau 

and Ridge/Blue Ridge 6 – 8% of the streams (ISA 3.2.4.2 and REA 4.2.2). Recreational fishing 

in lakes and streams is among the most popular outdoor recreational activities in the northeastern 

United States. Data from the 2006 National Survey of Fishing, Hunting, and Wildlife Associated 

Recreation (FHWAR) indicate that more than 9% of adults in this part of the country participate 

annually in freshwater fishing with 140 million freshwater fishing days. Based on studies 

conducted in the northeastern United States, Kaval and Loomis (2003) estimated average 

consumer surplus values per day of $35 for recreational fishing (in 2007 dollars). Therefore, the 

implied total annual value of freshwater fishing in the northeastern United States was $5 billion 

in 2006. We recognize that 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 given the magnitude of the resource, the length of time 

nitrogen and sulfur have been affecting freshwaters in the northeast and the level of monetary 

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damages calculated for the case study in the Adirondacks described in the next section we would 

expect these damages to be significant. 

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In general, inland surface waters such as lakes, rivers, and streams 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 United States 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 (Driscoll et al., 2001). 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. 

3.4.1.1 What is the value of the impaired recreational fishing and other cultural services? 

In the previous section we described the ecosystem services that are most likely to be 

affected by N and S deposition and summarized evidence regarding the current magnitude and 

values of recreational fishing services, the degree to which these services are impaired by 

existing NOy/SOx levels has not been quantified. To address this limitation, the REA 

(Appendix 8) provides insights into the magnitude of ecosystem service impairments. The 

REA provides quantitative estimates of selected ecosystem services impairments or 

enhancements for three main categories of ecosystem effects – aquatic acidification, terrestrial 

acidification, and aquatic nutrient enrichment6

. Within these three categories, the selection of 

specific ecosystem services for more in-depth analysis depended primarily on the expected 

magnitude of impairments and on the availability of appropriate data and modeling tools. 

The analysis of ecosystem service impairments due to aquatic acidification builds on the 

case study analysis of lakes in the New York Adirondacks. In this study estimates of changes in 

recreational fishing services are determined, as well as changes more broadly in “cultural” 

ecosystem services (including recreational, aesthetic, and nonuse services). First, the MAGIC 

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and models for this effect category. 

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model (REA, Appendix 8, Sec 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. These model runs assumed 

a 2010 “zero-out” emissions scenario (where all N and S deposition is eliminated) with a 

projected lag time between the elimination of emissions to observed improvement in ANC of 10 

years thus benefits results were calculated for the year 2020. These predictions were then 

extrapolated to the full universe of Adirondack lakes. Table 3-2 reports the number of 

“impacted” lakes in each year, where impact means that the lake is predicted to be below the 

ANC threshold under business-as-usual and above the threshold under pristine conditions. 

Table 3-2. Count of Impacted Lakes 

ANC Threshold 

(in μeq/L) Year Lake Count 

20 2005 0 

20 2020 107 

20 2050 95 

20 2100 74 

50 2005 0 

50 2020 244 

50 2050 222 

50 2100 200 

100 2005 0 

100 2020 430 

100 2050 404 

100 2100 354 

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Note: There are 1,076 lakes in the “Adirondack Region”. 

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. Table 3 – 3 

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summarizes the results and the present value of benefits and annualized benefits at 3 and 7% 

discount rates. 

Table 3-3. Present Value and Annualized Benefits, Adirondack Region 

Present Value Benefitsa

(in million of 2007 dollars) 

Annualized Benefitsb

(in million of 2007 dollars) ANC 

Threshold 

(in eq/L) 

3% Discount 

Rate 

7% Discount 

Rate 

3% Discount 

Rate 

7% Discount 

Rate 

20 $142.59 $60.05 $4.46 $3.94 

50 $285.15 $114.18 $8.91 $7.49 

100 $298.67 $120.61 $9.33 $7.91 

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a

 Annual benefits for 2010 to 2100 discounted to 2010. 

b Present value benefits annualized over 2009–2100. 

 Current annual impairments are most likely of a similar magnitude because, although 

current NOy/SOx levels are somewhat higher than those expected in 2020 (under business as 

usual – given expected emissions controls associated with Title IV regulations but no additional 

nitrogen or sulfur controls), and the affected NY population is also somewhat smaller (based on 

U.S. Census Bureau projections). 

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 WTP for future reductions (20% and 45%) 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. The survey that measured the benefits of 20% of the lakes improving 

indicated that terrestrial benefits were minor and econometric controls were used to adjust the 

willingness to pay estimate for those that suspected that terrestrial improvements were greater 

than described in the survey. The survey that measured the benefits of improving 45% of the 

total number of lakes also indicated that the benefits to forests and birds were significant. . 

The WTP estimates from the two versions of the survey were then (1) scaled to reflect 

predicted changes between business-as-usual and background conditions in 2020 (MAGIC lake 

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modeling results indicate that impaired lakes would decrease from 22 to 31% using background 

conditions with ANC increasing from 20 to 50μeq/L), and (2) aggregated across New York 

households. The scaling entails converting the average household willingness-to-pay for the 

improvements described in the Adirondacks surveys to an average household willingness-to-pay 

per percentage point of the total population of lakes improved.

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Table 3-4. The range of average household willingness to pay reflects the range in willingness to 

pay per percentage point of lakes improved described in the two versions of the survey. 

Estimates are provided at ANC 20, 50, and 100 to reflect the range of ANC discussed 

throughout the REA and this document and for consistency with the Random Utility Model 

analysis. 

Table 3-4. Aggregate Annual Benefit Estimates for the Zero-Out Scenario 

ANC 

 Threshold 

Reduction in 

Percentage of 

Unhealthy 

Lakes 

Range of Average 

Household WTP 

per Percentage 

Reduction 

Number of NY 

Households 

(in millions) 

Range of Annual 

Benefits 

(in millions of 2007 $) 

20 μeq/L 22% $2.63 $5.87 7.162 $410.6 $916.4 

50 μeq/L 31% $1.32 $3.76 7.162 $291.2 $829.4 

100 μeq/L 26% $2.63 $5.87 7.162 $491.6 $1,097.2 

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These results suggest that the value of avoiding current impairments to ecosystem 

services from Adirondack lakes are even higher than the estimate, because the estimates assume 

a lag of 10 years in which no benefits accrue and because the percent of impaired lakes is slightly 

higher today than expected in 2020 under business-as-usual. These results imply significant 

value to the public in addition to those derived from recreational fishing services. Note that the 

results are only applicable to improvements in the Adirondacks valued by residents of New 

York. If similar benefits exist in other acid-impacted areas, benefits for the nation as a whole 

could be substantial. The analysis provides results on only a subset of the impacts of acidification 

on ecosystem services and suggests that the overall impact on these services is likely to be 

substantial. 

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 Scaling is required because neither of the surveys administered by Banzhaf et al. (2006) describe improvements 

that correspond exactly to the improvement scenario modeled here. 

September, 2010 3-30 Do Not Quote or Cite 
1 3.4.2 Evidence for Adversity Related to Terrestrial Acidification 

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In the previous chapter of this document we discussed the effects of acidifying deposition 

on terrestrial ecosystems, especially forests. These include the observed decline and/or dieback 

in red spruce and sugar maple. These species are particularly sensitive to acidifying deposition 

and have ranges that overlap the areas of the U.S. where some of the highest levels of acidifying 

deposition occur. Additionally these species are present in the case study areas examined in the 

REA. As a result we chose to focus on red spruce and sugar maple as the species of interest for 

the analysis of ecosystem services presented in this section. 

A similar model to Figure 3-6 can be drawn for terrestrial acidification that links Bc:Al 

molar ratio to reduced tree growth to decreases in timber harvest, although we have less 

confidence in the significance of this linkage than we do for aquatic acidification. There are 

numerous services expected to be affected but the data and methods to adequately describe those 

losses does not as yet exist. These services include effects to forest health, water quality, and 

habitat, including decline in habitat for threatened and endangered species, decline in forest 

aesthetics, decline in forest productivity, increases in forest soil erosion and decreases in water 

retention (ISA, 2009; REA, 2009; Krieger, 2001). Forests in the northeastern United States 

provide several important and valuable provisioning 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 (NASS, 2008). Red spruce is also used in a variety of wood products 

and provides up to $100 million in economic value annually (USFS, 2006). 

Forests in the northeastern United States are also an important source of cultural 

ecosystem services, including nonuse (existence value for threatened and endangered species), 

recreational, and aesthetic services (ISA, 2009; REA, 2009). Red spruce forests are home to two 

federally listed species, the spruce-fir moss spider and the rock gnome lichen. The value of these 

two endangered species has not been estimated. 

Although we do not have the data to link acidification damages directly to economic 

values of lost recreational services in forests, these resources are valuable to the public. For 

example, most recent data from the National Survey on Recreation and the Environment (NSRE) 

indicate that, from 2004 to 2007, 31% of the U.S. adult (16 and older) population visited a 

September, 2010 3-31 Do Not Quote or Cite 
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wilderness or primitive area during the previous year, and 32% engaged in day hiking (Cordell et 

al., n.d.). A recent study suggests that the total annual value of off-road driving recreation was 

more than $9 billion, total and value of hunting and wildlife viewing was more than $4 billion 

each in the Northeastern United States in 2006 (Kaval and Loomis, 2003). Table 3-5 

summarizes data from the NSRE and the Fishing, Hunting, and Wildlife Related Activity Survey 

(U.S. DOI, 2007) along with average WTP estimates from Kaval and Loomis to estimate the 

total value of these services in the northeast. 

Table 3-5. Annual participation and value of outdoor (forest related) 

activity in the northeast. 

Recreational 

Activity 

Participation 

Rate 

(%) 

Activity 

Days 

(in 

Thousands) 

Avg. 

WTP Per 

Activity 

Day 

($2007) 

Total 

Value 

(in millions) 

Off Road Driving 16 366,336 $25.25 $9,250 

Hunting 5.5 83,821 $52.36 $4,380 

Wildlife Viewing 10 122,200 $34.46 $4,210 

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 In addition, fall color viewing is a recreational activity that is directly dependent on 

forest conditions. Sugar maple trees, in particular, are known for their bright colors and are, 

therefore, an essential aesthetic component of most fall color landscapes. Statistics on fall color 

viewing are much less available than for the other recreational and tourism activities; however, a 

few studies have documented the extent and significance of this activity. For example, Spencer 

and Holecek (2007) found that approximately 30% of residents in the Great Lakes area reported 

at least one trip in the previous year involving fall color viewing. In a separate study conducted 

in Vermont, Brown (2002) reported that more than 22% of households visiting Vermont in 2001 

made the trip primarily for the purpose of viewing fall colors. 

September, 2010 3-32 Do Not Quote or Cite 
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Two studies have estimated values for protecting high-elevation spruce forests in the 

Southern Appalachians. Kramer et al. (2003) conducted a contingent valuation study estimating 

households’ WTP for programs to protect remaining high-elevation spruce forests from damages 

associated with air pollution and insect infestation (Haefele et al., 1991; Holmes and Kramer, 

1995). The survey presented respondents with a sheet of color photographs representing three 

stages of forest decline and explained that, without forest protection programs, high-elevation 

spruce forests would all decline to worst conditions (with severe tree mortality) and two potential 

forest protection programs. Median household WTP was estimated to be roughly $29 (in 2007 

dollars) for the minimal program and $44 for the more extensive program. Another study by 

Jenkins, Sullivan, and Amacher (2002) estimated an aggregate annual value of $3.4 billion for 

avoiding a significant decline in the health of high-elevation spruce forests in the Southern 

Appalachian region. 

Table 3-6. Summary of Studies of Select Terrestrial Ecosystem Services 

Fall Color Viewing 30% Great Lakes area residents Spencer (2007) 

22% Vermont visitors Brown (2002) 

Protection of spruce Southern Appalachians 

$29 WTP per household for minimal program Kramer et al. (2003)

$44 WTP per household for extensive program

$3.4 b Aggregate annual value Jenkins (2002) 

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Forests in the northeastern United States also support and provide a wide variety of 

valuable regulating services, including soil stabilization and erosion control, water regulation, 

and climate regulation (Krieger, 2001). Forest vegetation plays an important role in maintaining 

soils in order to reduce erosion, runoff, and sedimentation that can adversely impact surface 

waters. In addition to protecting the quality of water in this way, forests also help store and help 

regulate the quantities and temporal discharge patterns of water in watersheds. Forests also play 

September, 2010 3-33 Do Not Quote or Cite 
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an important role in carbon sequestration at both regional and global scales. The total value of 

these ecosystem services is very difficult to quantify. 

3.4.2.1 What is the value of current ecosystem service impairments? The REA 

Appendix 8 describes an analysis of ecosystem service impairments associated with the impacts 

of terrestrial acidification on the forest product provisioning services from two commercially 

important tree species on unmanaged forests – sugar maple and red spruce - that are particularly 

sensitive to the effects of acidification. Evidence of effects due to terrestrial acidification is 

particularly strong for these two species whose range includes the northeastern U.S. where levels 

of nitrogen and sulfur deposition have historically been relatively high, however more 

widespread impacts that include other tree species are also possible. We acknowledge that there 

may be some beneficial fertilization effects of nitrogen deposition however given the complexity 

of the nitrogen cycle it is not possible to quantify all those effects here. There is a detailed 

discussion of nitrogen fertilization effects in Chapter 4. 

In an exploratory study that is still under development we used data from the USFS 

Forest Inventory and Analysis (FIA) database, to estimate an exposure-response relationship for 

each species to measure the average negative effect of critical load exceedances (CLEs) of 

nitrogen and sulfur deposition on annual tree growth. These estimated relationships were then 

applied to sugar maple and red spruce stocks in the Northeast and North central regions to 

estimate the average percent increase in annual tree growth that would occur if all CLEs were 

eliminated. To estimate the aggregate-level forest market impacts of eliminating CLEs starting 

in the year 2000, the tree-level growth adjustments were applied using the Forest and 

Agricultural Sector Optimization Model (FASOM), which is a dynamic optimization model of 

the U.S. forest and agricultural sectors. The model results are reported as the present discounted 

values of future welfare changes in the forestry sector (in 5-year increments from 2000 to 2080) 

due to increased tree growth. Summing over this 80-year period, the total present value of these 

welfare gains is $40.705 million (in 2006 dollars, using a 4% discount rate). On an annualized 

basis (at 4%), this is equivalent to $1.64 million per year. These estimates can also be 

interpreted as the current value of impairments to forest provisioning services provided by red 

spruce and sugar maple due to acidification effects from nitrogen and sulfur. These results 

should be considered very uncertain due to the pending revision of the exposure – response curve 

and release of an updated version of the FASOM model. 

September, 2010 3-34 Do Not Quote or Cite 
1 3.4.3 Evidence for Adversity Related to Aquatic Nutrient Enrichment 

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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 (U.S. EPA, 2009). 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. It is not known, however, what percentage of this value is directly attributable to or 

dependent upon the estuaries in these states. Based on commercial landings in Maryland and 

Virginia, the values for three key species—blue crab, striped bass, and menhaden- totaled nearly 

$69 million in 2007 in the Chesapeake Bay alone. 

Assessing how eutrophication in estuaries affects fishery resources requires bioeconomic 

models (i.e., models that combine biological models of fish population dynamics with economic 

models describing fish harvesting and consumption decisions), but relatively few exist (Knowler, 

2002). Kahn and Kemp (1985) estimated that a 50% decline in submerged aquatic vegetation 

(SAV) from levels existing in the late 1970s (similar to current levels [Chesapeake Bay Program, 

2008]) would decrease the net social benefits from striped bass by $16 million (in 2007 dollars). 

In a separate analysis, Anderson (1989) modeled blue crab harvests under baseline conditions 

and under conditions with “full restoration” of SAV. In equilibrium, the increase in annual 

producer surplus and consumer surplus with full restoration of SAV was estimated to be $7.9 

million (in 2007 dollars) or an 11% increase from current service provision from blue crab alone. 

Mistiaen et al. (2003) found that reductions in dissolved oxygen (DO) cause a statistically 

significant reduction in commercial harvest and revenues crab harvests. For the Patuxent River 

alone, a simulated reduction of DO from 5.6 to 4.0 mg/L was estimated to reduce crab harvests 

by 49% and reduce total annual earnings in the fishery by $275,000 (in 2007 dollars). While 

these values do not quantify the increase in terms of atmospheric loadings alone, the estimated 

20% loading to the Potomac River watershed (REA 5.2.4) from atmospheric deposition indicates 

that the benefits apportioned to deposition are significant. 

In addition, eutrophication in estuaries may also affect the demand for seafood. For 

example, a well-publicized toxic pfiesteria bloom in the Maryland Eastern Shore in 1997 led to 

an estimated $56 million (in 2007 dollars) in lost seafood sales for 360 seafood firms in 

September, 2010 3-35 Do Not Quote or Cite 
Maryland in the months following the outbreak (Lipton, 1999). Surveys by Whitehead, Haab, 

and Parsons (2003) and Parsons et al. (2006) indicated a reduction in consumer surplus due to 

eutrophication-related fish kills ranging from $2 to $5 per seafood meal.

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estimated aggregate consumer surplus losses of $43 million to $84 million (in 2007 dollars) in 

the month after a fish kill. 

As mentioned in the REA (5.2.1.3), estuaries in the eastern United States also provide an 

important and substantial variety of cultural ecosystem services, including water-based 

recreational and aesthetic services. For example, FHWAR data indicate that 4.8% 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 (U.S. DOI, 2007). Based on 

estimates in Section 5.2.1.3 of the REA, total recreational consumer surplus value from these 

saltwater fishing days was approximately $1.3 billion (in 2007 dollars). Recreational 

participation estimates for several other coastal recreational activities are also available for 

1999–2000 from the NSRE. Almost 6 million individuals participated in motorboating in coastal 

states from North Carolina to Massachusetts. Again, based on analysis in the REA, the aggregate 

value of these coastal motorboating outings was $2 billion per year. Almost 7 million people 

participated in birdwatching, for a total of almost 175 million days per year, and more than 3 

million participated in visits to nonbeach coastal waterside areas, for a total of more than 35 

million days per year. 

Estuaries and marshes have the potential to support a wide range of regulating services, 

including climate, biological, and water regulation; pollution detoxification; erosion prevention; 

and protection against natural hazards (MEA, 2005c). The relative lack of empirical models and 

valuation studies imposes obstacles to the estimation of ecosystem services affected by nitrogen 

deposition. While atmospheric deposition contributes to eutrophication there is uncertainty in 

separating the effects of atmospheric nitrogen from nitrogen reaching the estuaries from many 

other sources. 

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different types of information included in the survey. 

September, 2010 3-36 Do Not Quote or Cite 
1 Table 3-7. Summary of Values for Current Levels of Services and Changes in 

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Service Levels in $2007 

Ecosystem Service Area or Population 

Affected 

Value 

($2007) 

Species 

Total Catch – 

Commercial Fishing 

14 east coast states $1.5 b/yr 

MD/VA $69 m/yr blue crab, striped bass, 

menhaden 

Change in Ecosystem 

Service 

Value of 

Change 

($2007) 

50% decline in SAV Chesapeake Bay 

 $16 m/yr striped bass 

Full restoration SAV Chesapeake Bay 

 $ 8 m/yr blue crab 

0.4% mg/L decrease DO Patuxent River 

 $275 th/yr 

 

 49% blue crab harvest 

HAB 1997 MD eastern 

shore $56 m loss to seafood industry 

 

 $43-84 m sustained loss over months 

Ecosystem Service Participation 

Days 

Value ($2007) 

Saltwater fishing 4.6% pop. MA-NC 26m days $1.3 b/yr 

Motorboating 6 million 2 b/yr 

Bird watching 7 million 175 m days 

Non-beach coastal visits 3 million 35 m days 

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3.4.3.1 Value of aquatic ecosystem service impairments from current levels nutrient 

enrichment 

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The aquatic nutrient enrichment case study relied on the NOAA Eutrophication Index as 

the indicator, which includes dissolved oxygen, HABs, loss of SAV and loss of water clarity. 

There are methods available to link some of the components to ecosystem services, most notably 

loss of SAV and reductions in DO. The REA analysis estimates the change in several ecosystem 

services including recreational fishing, boating, beach use, aesthetic services and nonuse 

services. The REA focuses on two major East Coast estuaries – the Chesapeake Bay and the 

Neuse River. Both estuaries receive between 20%-30% percent of their annual nitrogen loadings 

through atmospheric deposition and both are showing symptoms of eutrophication. The analysis 

uses and adapts results from several existing studies to approximate effects on several ecosystem 

services, including commercial fishing, recreation, aesthetic enjoyment, and nonuse values. For 

example, it is estimated that atmospheric nitrogen decreases the annual benefits of recreational 

fishing, boating, and beach use in the Chesapeake Bay by $43-$217 million, $3-8 million, and 

$124 million respectively, and reduces annual aesthetic benefits to nearshore residents by $39-

102 million. In the Neuse River, the value of annual commercial crab fishing services would be 

between $0.1-1 million higher without the contribution of atmospheric nitrogen, and recreation 

fishing services in the larger Albermarle Pamlico Sound estuary system (which includes the 

Neuse) would be $1-8 million greater per year. 

Table 3-8. Summary of Annual Damages to Services due to Atmospheric Loading 

Ecosystem Service Annual Value ($2007) Waterbody Affected 

Recreational Saltwater Fishing $43-217 b Chesapeake Bay 

$1-8 m Albemarle Pamlico Sound

Beach Use $39-102 m Chesapeake Bay 

Boating $3-8 m Chesapeake Bay 

Commercial Crab Fishing $0.1-1 m Neuse River 

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September, 2010 3-38 Do Not Quote or Cite 
1 3.4.4 Evidence for Adversity Related to Terrestrial Nutrient Enrichment 

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The ecosystem service impacts of terrestrial nutrient enrichment in unmanaged ecosystems 

include primarily cultural and regulating services. In CSS areas, concerns focus on a decline in 

CSS and an increase in nonnative grasses and other species, impacts on the viability of threatened 

and endangered species associated with CSS, and an increase in fire frequency. Changes in MCF 

include changes in habitat suitability and increased tree mortality, increased fire intensity, and a 

change in the forest’s nutrient cycling that may affect surface water quality through nitrate 

leaching (EPA, 2008). 

The terrestrial nutrient enrichment case study relies on benchmark deposition levels for 

various species and ecosystems as indicators of ecosystem response. While it would be expected 

that deposition above those levels would have deleterious effects on the provision of ecosystem 

services in those areas, at this time it is possible only to describe the magnitude of the some of 

the services currently being provided. Methods are not yet available to allow estimation of 

changes in services due to nitrogen deposition. 

The value that California residents and the U.S. population as a whole place on CSS and 

MCF habitats is reflected in the various federal, state, and local government measures that have 

been put in place to protect these habitats. Threatened and endangered species are protected by 

the Endangered Species Act. The State of California passed the Natural Communities 

Conservation Planning Program (NCCP) in 1991, and CSS was the first habitat identified for 

protection under the program (seewww.dfg.ca.gov/habcon/nccp). It is estimated that only 10 – 

15% of the original extent of CSS habitat remains (NPS.gov/cabr/naturescience/coastal-sagescrub-and-southern-chaparrel-communities.htm). Private organizations such as The Nature 

Conservancy, the Audubon Society, and local land trusts also protect and restore CSS and MCF 

habitat. 

CSS and MCF are found in numerous recreation areas in California. Three national parks 

and monuments in California contain CSS, including Cabrillo National Monument, Channel 

Islands National Park, and Santa Monica National Recreation Area. All three parks showcase 

CSS habitat with educational programs and information provided to visitors, guided hikes, and 

research projects focused on understanding and preserving CSS. Over a million visitors traveled 

through these three parks in 2008. MCF is highlighted in Sequoia and Kings Canyon National 

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Park, Yosemite National Park, and Lassen Volcanic National Park, where more than 5 million 

people visited in 2008. 

The 2006 FHWAR for California (DOI, 2007) reports on the number of individuals 

involved in fishing, hunting, and wildlife viewing in California. Millions of people are involved 

in just these three activities each year. The quality of these trips depends in part on the health of 

the ecosystems and their ability to support the diversity of plants and animals found in important 

habitats found in CSS or MCF ecosystems and the parks associated with those ecosystems. 

Based on analyses in Section 5.3.1.3 of the REA (U.S.EPA, 2009), average values of the total 

benefits in 2006 from fishing, hunting, and wildlife viewing away from home in California were 

approximately $947 million, $169 million, and $3.59 billion, respectively. In addition, data from 

California State Parks (2003) indicate that in 2002, 68.7% of adult residents participated in trail 

hiking for an average of 24.1 days per year. The analyses in the REA (U.S.EPA, 2009) indicate 

that the aggregate annual benefit for California residents from trail hiking in 2007 was $11.59 

billion. It is not currently possible to quantify the loss in value of services due to nitrogen 

deposition as those losses are already reflected in the estimates of the contemporaneous total vale 

of these recreational activities. Restoration of services through decreases in nitrogen deposition 

would likely increase the total value of recreational services. 

Table 3-9. Summary of Current Levels of Ecosystem Services 

Activity Participation # of 

days/yr

Average 

WTP 

Annual Aggregate 

Value ($2007 in 

millions) 

Trail Hiking 68.7% of CA 

population 

453 m $25.59 115,900 

Fishing 1.7 m 19 m $48.86 947 

Wildlife 

Viewing 

6.2 m 45 m $79.81 3,600 

Hunting 0.28 m 3.3 m $50.10 169 

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Sources: 2006 FHWAR for California (DOI, 2007), California State Parks (2003), 

Kaval and Loomis(2003) 

September, 2010 3-40 Do Not Quote or Cite 
CSS and MCF are home to a number of important and rare species and habitat types. CSS 

displays richness in biodiversity with more than 550 herbaceous annual and perennial species. Of 

these herbs, nearly half are endangered, sensitive, or of special status (Burger et al., 2003). 

Additionally, avian, arthropod, herpetofauna, and mammalian species live in CSS habitat or use 

the habitat for breeding or foraging. Communities of CSS are home to three important federally 

endangered species: the Quino checkerspot butterfly, the kangaroo rat and the California 

gnatcatcher. MCF is home to one federally endangered species (mountain yellow-legged frog) 

and a number of state-level sensitive species. The Audubon Society lists 28 important bird areas 

in CSS habitat and at least 5 in MCF in California (

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The terrestrial enrichment case study in Section 5.3.1.3 of the REA and Section 3.3.5 of 

the ISA identified fire regulation as a service that could be affected by nutrient 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 (National Association of State Foresters [NASF], 2009). It is not possible at this time to 

quantify the contribution of nitrogen deposition, among many other factors, to increased fire risk. 

The CSS and MCF were selected as case studies for terrestrial enrichment because of the 

potential that these areas could be adversely affected by excessive N deposition. To date, the 

detailed studies needed to identify the magnitude of the adverse impacts due to N deposition 

have not been completed. Based on available data, this report provides a qualitative discussion of 

the services offered by CSS and MCF and a sense of the scale of benefits associated with these 

services. California is famous for its recreational opportunities and beautiful landscapes. CSS 

and MCF are an integral part of the California landscape, and together the ranges of these 

habitats include the densely populated and valuable coastline and the mountain areas. Through 

recreation and scenic value, these habitats affect the lives of millions of California residents and 

tourists. Numerous threatened and endangered species at both the state and federal levels reside 

in CSS and MCF. Both habitats may play an important role in wildfire frequency and intensity, 

an extremely important problem for California. The potentially high value of the ecosystem 

services provided by CSS and MCF justify careful attention to the long-term viability of these 

habitats. 

9

 Important Bird Areas are sites that provide essential habitat for one or more species of bird. 

September, 2010 3-41 Do Not Quote or Cite 
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2 4 ADDRESSING THE ADEQUACY OF THE CURRENT STANDARDS 

 Based on the information in Chapters 2 and 3, we conclude that there is support in the 

available effects-based evidence for consideration of secondary standards for NOX and SOX that 

are protective against adverse ecological effects associated with deposition of NOX and SOX to 

sensitive ecosystems. Having reached this general conclusion, we then to the extent possible 

evaluate the adequacy of the current NOX and SOX secondary standards by considering to what 

degree risks to sensitive ecosystems would be expected to occur in areas that meet the current 

standards. Staff 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, staff have taken into account the uncertainties and limitations in the scientific 

evidence. This chapter addresses key policy relevant questions that inform our determination 

regarding the adequacy of the structure and levels of the current secondary standards. The 

chapter begins with a discussion of the structure of the current standards, followed by a 

presentation of information on recent air quality relative to the existing standards, recent NOX

and SOX deposition levels, evaluation of recent deposition levels relative to levels where adverse 

ecological effects have been observed, and a set of conclusions regarding the adequacy of the 

current structure and levels of the standards. Acidification occurs over extended periods and 

the ability of both terrestrial and aquatic systems to recover is dependent upon not only the 

decrease in acidic deposition, but the ability of these ecosystems to generate cations needed for 

nutrients and base cation supply. As a result, given the same decrease in deposition, ecosystems 

with high levels of base cation replacement will recover faster than those with low levels. 

4.1 Are The Structures Of The Current NOx And SOx Secondary Standards Based On 

Relevant Ecological Indicators Such That They Are Adequate To Determine And 

Protect Public Welfare Against Adverse Effects On Ecosystems? 

The current secondary NOX and SOX standards are intended to protect against adverse 

effects to public welfare. For NOX, the current secondary standard was set identical to the 

September, 2010 Do Not Quote or Cite 4-1
primary standard1

, e.g. an annual standard set for NO2 to protect against adverse effects on 

vegetation from direct exposure to ambient NOX. For SOX, 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 in this review 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 SOx and NOx. There is sufficient evidence to suggest that 

the levels of the current standards are likely adequate to protect against direct phytotoxic effects. 

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The ISA however, has established that the major effects of concern for this review of the 

NOX and SOX standards are associated with deposition of N and S caused by atmospheric 

concentrations of NOX and SOX (see Chapter 2). As discussed in the following sections, 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 N and S deposition. Thus, by using atmospheric NO2 and SO2 concentrations as 

indicators, the current standards address only a fraction of total atmospheric NOX and SOX, and 

do not take into account the effects from deposition of total atmospheric NOX and SOX. By 

addressing short-term concentrations, the current SO2 standard, while protective against direct 

foliar effects from gaseous SOX, does not take into account the findings of effects in the ISA, 

which notes the relationship between annual deposition of S 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 which occurs over time, as 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 standard. 

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annual distribution of the 1 hour daily maximum of the concentration of NO2. The current secondary standard 

remains as it was set in 1971. 

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The levels of the current standards also are not well suited to dealing with depositionbased effects of NOx and SOx. Current standards are specified as allowable single atmospheric 

concentration levels for NO2 or SO2. This type of structure does not take into account variability 

in the atmospheric and ecological factors that may alter the effects of NOX and SOX on public 

welfare. Consistent with section 108 of the CAA, the ISA includes in the air quality criteria 

consideration of how these variable factors impact the effects of ambient NOX and SOX on public 

welfare. See CAA section 108 (a)(2)(A) requiring air quality criteria to include information on 

“those variable factors (including atmospheric conditions) which of themselves or in 

combination with other factors may alter the effects on ... welfare of such air pollutant”. 

Secondary standards are intended to address a wide variety of effects occurring in different types 

of environments and ecosystems. Ecosystems are not uniformly distributed either spatially or 

temporally in their sensitivity to air pollution. Therefore, failure to account for the major 

determinants of variability, including geological and soil characteristics related to the sensitivity 

to acidification 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. Finally, given the mismatch of all of the other elements of the current secondary 

NAAQS with deposition-based effects, the form of those standards will also be mismatched. 

Because most areas of the U.S. are in attainment with the current NO2 and SOX standards, 

it is possible to evaluate current conditions, and evaluate the impact on public welfare from the 

current effects on ecosystems from NOX and SOX deposition in areas that attain the current 

standards that use NO2 and SO2 as indicators. In addition, this chapter qualitatively addresses 

the adequacy of the structures of the existing standards relative to ecologically relevant standards 

for NOX and SOX, and sets up arguments for developing an ecologically relevant structure for the 

standards as described in Chapter 5. 

4.2 To What Extent Are The Structures Of The Current NOX AND SOX Secondary 

Standards Meaningfully Related To Relevant Ecological Indicators Of Public 

Welfare Effects? 

The current secondary standard for NOX, set in 1971, using NO2 as the atmospheric 

indicator, is 0.053 parts per million (ppm) (100 micrograms per cubic meter of air [μg/m3]), 

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

This standard was selected to provide protection to the public welfare against acute injury to 

vegetation from direct exposure and resulting phytoxicity. During the last review of the NOX

standards, impacts associated with chronic acidification and eutrophication from NOx deposition 

were acknowledged, but the relationships between atmospheric concentrations of NOx and levels 

of acidification and eutrophication and associated welfare impacts were determined to be too 

uncertain to be useful as a basis for setting a national secondary standard (USEPA 1995). 

The current secondary standard for SOX, set in 1971, uses SO2 as the atmospheric 

indicator, is a 3-hour average of 0.5 ppm, not to be exceeded more than once per year. This 

standard was selected to provide protection to the public welfare against acute injury to 

vegetation. In the last review of the SOX secondary standard, impacts associated with chronic 

acidification were acknowledged, but the relationships between atmospheric concentrations of 

SOx and levels of acidification, along with the complex interactions between SOX and NOX in 

acidification processes, were cited as critical uncertainties which made the setting of secondary 

NAAQS to protect against acidification inappropriate at that time (USEPA 1982). 

In the previous separate reviews of the NOX and SOX secondary standards, EPA 

acknowledged in each review the additional impacts of NOX and SOX on public welfare through 

the longer term impact of the pollutants once deposited to ecosystems. However, the previous 

reviews cited numerous uncertainties as the basis for not directly addressing those impacts in the 

setting of secondary standards. In addition, these previous reviews did not consider the common 

pathways of impact for both nitrogen and sulfur acting on the same ecosystem endpoints. 

Three issues arise that call into question the ecological relevance of the current structure 

of the secondary standards for NOX and SOX. One issue is the 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 NO2 and SO2 as measured by the current 

standards. Even though the NO2 standard uses an annual average of NO2, it is focused on the 

annual average of 1-hour NO2 concentrations, rather than on a cumulative metric or an averaging 

metric based on daily or monthly averages. A second issue is the choice of atmospheric 

indicators. NO2 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 

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deposition of NOX and SOX 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

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 total sulfur deposition, and thus all 

forms of oxidized nitrogen and oxidized sulfur that are deposited will contribute to effects on 

ecosystems. This suggests that more comprehensive atmospheric indicators should be 

considered in designing ecologically relevant standards. Further discussions of the need for 

more ecologically relevant atmospheric indicators as well as the relative contributions to 

deposition from various species of NOX and SOX can be in found in Chapters 5 and 6 below. 

The third issue is that the current standards reflect separate assessments of the two individual 

pollutants, NO2 and SO2, rather than assessing the joint impacts of deposition of NOX and SOX to 

ecosystems, recognizing the role that each pollutant plays in jointly affecting ecosystem 

indicators, functions, and services. The clearest example of this interaction is in assessment of 

the impacts of acidifying deposition on aquatic ecosystems. 

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Acidification in an aquatic ecosystem depends on the total acidifying potential of the 

deposition of both N and S from both atmospheric deposition of NOX and SOX as well as the 

inputs from other sources of N and S 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 NOX and SOX should be a critical element of 

the conceptual framework for ecologically relevant standards. There are also important 

interactions between NOX and SOX and reduced forms of nitrogen, which also contribute to 

acidification and nutrient enrichment. Although the standards do not directly address reduced 

forms of nitrogen in the atmosphere, e.g. they do not require specific levels of reduced nitrogen, 

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 NOX and SOX. 

Consideration will also have to be given to account for loadings coming from non-atmospheric 

sources as ecosystems will respond to these sources as well. 

In addition to the fundamental issues discussed above, the current structures of the 

standards do not address the complexities in the responses of ecosystems to deposition of NOX

and SOX. Ecosystems contain complex groupings of organisms that respond in various ways to 

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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 than 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 which control rates of deposition as well 

as ecological factors which 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. 

12 4.3 To What Extent Do Current Monitoring Networks Provide A Sufficient Basis For 

Determining The Adequacy of Current Secondary NOX and SOX Standards? 

Staff have closely evaluated whether levels of N and S allowed by current standards are 

requisite to protect public welfare. Doing so requires relating atmospheric concentrations with 

deposition, exposure pathways, and measured effects on ecologic receptors. A combination of 

monitoring and air quality model applications is useful in linking atmospheric concentrations to 

ecological effects. There are over 1000 ground level monitoring platforms (Figures 4-1 and 4-2 

and Table 4 -1) that provide measurements of some form of atmospheric nitrogen or sulfur. The 

key pollutants for this assessment are total oxidized nitrogen (NOY), total reduced nitrogen 

(NHX), and total oxidized sulfur which is referenced herein as (SOX) and defined as the sum of 

SO2 (gas) and particulate sulfate.. Total reactive oxidized atmospheric nitrogen, NOY, is defined 

as NOX (NO and NO2) and all oxidized NOX products: NOy = NO2 + NO + HNO3 + PAN 

+2N2O5 + HONO+ NO3 + organic nitrates + particulate NO3 (Finlayson-Pitts and Pitts, 2000). 

This definition of NOY reflects the operational principles of standard measurement techniques in 

which all oxidized nitrogen species are converted to nitrogen oxide (NO) through catalytic 

reduction and the resulting NO is detected through luminescence. Thus, NOY is truly defined as 

total oxidized nitrogen as converted to NO, essentially representing all oxidized nitrogen atoms. 

NOY is not a strict representation of the all moles of oxidized nitrogen as the diatomic 

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1 Table 4-1. Summary of Monitoring Networks.

Network Number 

of Sites 

Species Measured Sampling 

Frequenc

y 

Comments 

All Sulfur Sites 

NCore 82 SO2 Hourly Includes 20 rural sites 

SEARCH 8 SO2 Hourly Includes 3 rural sites 

SO2 751 SO2 Hourly NAMS/SLAMS/PAM

S for 2008 

PM Speciation 242 Sulfates 24-hour Measurements of 

Sulfates (88403) 

identified in AQS for 

Trends and 

Supplemental 

Speciation monitoring 

type for 2008 

IMPROVE 215 Sulfates 24-hour IMPROVE 

Monitoring Sites with 

Measurements of 

Sulfates (88403) 

identified in AQS 

CASTNET 88 Sulfates Weekly 

Ave. 

EPA & NPS 

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1 Table 4-1 Summary of Monitoring Networks (continued) 

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Network Number of 

Sites 

Species 

Measured 

Sampling 

Frequency 

Comments 

All Nitrogen Sites 

NCore 82 NO/NOy Hourly Includes 20 rural sites 

SEARCH 8 NO/NO2/NOy/HNO

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Hourly Includes 3 rural sites 

PAMS 119 NO2/NOx Hourly Official sites as of 

12/09 

SLAMS 643 NO/NO2/NOx/NOy Hourly All SLAMS 

Monitoring Sites with 

Measurements of NO, 

NO2, NOX or NOY in 

2009 identified in 

AQS 

NOY 59 NOy Hourly All Monitoring Sites 

with Measurements of 

NOY in 2009 

identified in AQS, 

regardless of 

Monitoring Type 

IMPROVE 214 Nitrates 24-hour MPROVE Monitoring 

Sites with 

Measurements of 

Nitrates (88306) 

identified in AQS 

CASTNET 88 Nitrates Weekly 

Ave. 

EPA & NPS 

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Figure 4-1 Routinely operating surface monitoring stations measuring forms of 

 atmospheric nitrogen. 

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Figure 4-2 Routinely operating surface monitoring stations measuring forms of 

atmospheric sulfur. All site locations measure both SO2 and sulfate except for 

the green SO2 only sites. 

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nitrogen species such as N2O5 yield 2 moles of NO. This definition is consistent with the 

relationship between atmospheric nitrogen and acidification processes as the reported NOY

provides a direct estimate of the potential equivalents available for acidification. Total reduced 

nitrogen (NHX) includes ammonia, NH3, plus ammonium, NH4 (EPA, 2008). Reduced nitrogen 

plus oxidized nitrogen is referred to as total reactive nitrogen. Total oxidized sulfur (SOX) 

includes SO2 gas and particulate sulfate, SO4. These species are converted to mass of sulfur 

which is used directly, or converted to charge equivalents, in deposition analyses linking 

atmospheric deposition and ecosystem models. Ammonium and sulfate are components of 

atmospheric particulate matter as well as directly measured and modeled in precipitation as 

direct deposition components. As discussed in this section, there are only very limited routine 

measurements of total oxidized and reduced nitrogen. In addition, existing monitoring networks 

do not provide adequate geographic coverage to fully assess concentrations and deposition of 

reactive nitrogen and sulfur in and near sensitive ecosystems. 

 The principal monitoring networks include the regulatory based State and Local Air 

Monitoring Stations (SLAMS) providing mostly urban-based SO2, NO and NOX, the PM2.5

chemical speciation networks Interagency Monitoring of Protected Visual Environments 

(IMPROVE) and EPA’s Chemical Speciation Network (CSN) providing particle bound sulfate 

and nitrate, and the Clean Air Status and Trends Network (CASTNET) providing weekly 

averaged values of SO2, nitric acid, and particle bound sulfate, nitrate and ammonium. The 

private sector supported South Eastern Aerosol Research and Characterization (SEARCH) Study 

network of 4-8 sites in the Southeast provides the only routinely operating source of true 

continuous NO2, ammonia, and nitric acid measurements. SEARCH also provides PM2.5 size 

fractions of nitrate and sulfate. Collectively, the SLAMS, Photochemical Assessment 

Measurement Stations (PAMS), SEARCH and NCore networks will provide over 100 sites 

measuring NOY (Figure 4-3). The NCore network (Scheffe et al., 2009) is a multiple pollutant 

network with co-located measurements of key trace gases (CO, SO2, O3, NO and NOY), PM2.5 

and PM(10-2.5) mass and PM2.5 chemical speciation. Additional air pollutants, particularly volatile 

organic compounds (VOCs), will be measured at those sites that are part of the existing PAMS 

and National Air Toxics Trends (NATTS) platforms. The NATTS (EPA, 2008) include 27 

stations across the U.S. that monitor for a variety of hazardous air pollutants and are intended to 

remain in place to provide a long-term record. Additional 

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Figure 4-3. Anticipated network of surface based NOY stations based on 2009 network 

design plans. The NCore stations are scheduled to be operating by January, 

2011. 

measurements of ammonia and possibly true NO2 are under consideration. True NO2 is noted to 

differentiate from the NO2 determined through routine regulatory networks that have known 

variable positive bias for NO2. The network currently is being deployed and expected to be 

operational with nearly 75 sites by January, 2011. The sites are intended to serve as central site 

monitors capturing broadly representative (e.g., not strongly influenced by nearby sources) air 

quality in a suite of major and mid size cities, and approximately 20 sites are located in rural 

locations. 

There are significant measurement gaps for characterizing NOY, NHX and SO2 in the 

nations ambient air observation networks (EPA, 2008) that lead to greater reliance on air quality 

modeling simulations to describe current conditions. National design of routinely operating 

ambient air monitoring networks is driven mostly by data uses associated with implementing 

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primary NAAQS, with noted exceptions of the CASTNET and IMPROVE networks. In 

addition to significant spatial gaps in sensitive ecosystem areas that arise from a population 

oriented network design, the current measurements for primary and secondary nitrogen are 

markedly different and in some instances of negligible value for secondary NOX and SOX 

standards. For example, a true NOX (NO plus NO2) measurement typically would capture less 

than 50% (see discussion below) of the total regional NOY mass in rural locations as the more 

aged air masses contain significant oxidized nitrogen products in addition to NOX. Note that the 

NOX monitors used for NAAQS primary compliance purposes capture varying amounts of 

transformed nitrogen species; however, the method provides biased low estimates with 

significant airshed induced variability relative to true NOY. With the exception of the SEARCH 

network in the Southeast, there are virtually no routine networks that measure ammonia, 

although EPA is considering options for ammonia sampling in CASTNET and NCORE 

networks. Ammonium is reported in EPA chemical speciation networks, although the values are 

believed to be biased low due to ammonia volatization. 

CASTNET provides mostly rural measurements of SO2, total nitrate, and ammonium, and 

affords an existing infrastructure useful for future monitoring in support of a potential NOx and 

SOx secondary standard. However, the lack of NOY, SOX and NHX measurements in sensitive 

ecosystems will require attention in the N/S secondary standard proposal. 

As a result of the limited monitoring networks for NOY and SOX in sensitive ecosystems, 

we are unable to use current ambient monitoring data to adequately link measured current 

atmospheric concentrations to ecological effects transmitted through deposition. At this time for 

the purpose of illustrating current atmospheric conditions, we supplement the available 

monitoring data with the use of sophisticated atmospheric modeling conducted using EPA’s 

CMAQ model (as discussed in Chapter 7). 

4.3.1 To what extent does the NADP monitoring network provide an adequate 

characterization of deposition and what are the major limitations? 

The National Atmospheric Deposition Program (NADP) includes approximately 250 sites 

(Figure 4-4) across the U.S. providing annual total wet deposition based on weekly averaged 

measures of wet deposition of nitrate, ammonium, sulfate and other ions based on the 

concentrations of these ions in precipitation samples. Meteorological models have difficulty in 

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capturing the correct spatial and temporal features of precipitation events, raising the importance 

of the NADP as a principal source of precipitation chemistry. The NADP has enabled several 

organizations to participate in a measurement program with a centralized laboratory affording 

measurement and analysis protocol consistency nationwide. Virtually every CASTNET site is 

located at an NADP site and the combined NADP/CASTNET infrastructure is a starting point for 

discussions addressing future NOX and SOX monitoring needs. The organic bound nitrogen is 

not analyzed routinely in NADP samples. Consideration might be given to adding NADP sites 

in locations where ambient air monitoring is conducted to assess compliance with a secondary 

NOX/ SOX standard. 

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Figure 4-4. Location of approximately 250 National Atmospheric Deposition Monitoring 

(NADP) National Trends Network (NTN) sites illustrating annual 

ammonium deposition for 2005. Weekly values of precipitation based nitrate, 

sulfate and ammonium are provided by NADP. 

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4.3.2 What are the relative strengths and important gaps in existing atmospheric 

monitoring networks to address a combined NOX/SOX secondary standard? 

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Of the currently operating monitoring networks, precipitation based sulfate, ammonium 

and nitrate measurements provided by the NADP are the most relevant measurements that would 

support the secondary standard as they provide atmospheric deposition inputs that drive 

ecosystem models, and NADP site locations generally include acid sensitive areas. However, 

there are significant gaps in ambient air (aerosols and gases) monitoring networks for the 

measurement of the likely ambient indicators of NOY, SO2, and SO4. CASTNET filter packs 

provide the most relevant source of ambient sulfate (SO4) measurements as the open inlet of the 

filter packs incorporates the full range of particle sizes that contribute to deposition. The SO2

measurements from CASTNET represent about 10% of all SO2 sites nationally, but are 

especially relevant based on their locations in rural and regional settings, although CASTNET is 

not as spatially extensive (breadth and resolution) as the NADP network of precipitation sites. 

Although CASTNET does provide measurements of total ambient nitrate, other oxidized 

nitrogen species constituting a more complete NOY budget are not captured. In their current 

configuration, the State and local monitoring networks virtually offer no support for a secondary 

NOX/ SOX standard due to their urban based site orientation and exclusion of important oxidized 

nitrogen species (e.g., nitrates and PAN). The chemical speciation networks, including rural 

based IMPROVE, all provide ambient sulfate measurements based on on a 2.5μ size cut. While 

the sulfate mass within that size fraction may constitute 80% or greater of the ambient sulfate 

budget, the missing larger size particles can contribute significantly to sulfate deposition due to 

their relatively high gravitationally driven deposition velocities. Finally, there are virtually no 

ambient ammonia measurements routinely collected in acid sensitive areas. CASTNET does 

provide ammonium measurements, but the routine speciation networks that report ammonium 

have expected artifacts due to ammonia offgassing from nylon filters. 

 Although this summary of existing networks suggests significant challenges in meeting 

the monitoring needs of a new NOX/ SOX standard, the networks do serve as a useful building 

block for moving forward. The site locations of NADP and CASTNET offer an infrastructure 

to accommodate additional instruments. The NCORE network has introduced nearly 75 NOY

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trace level SO2 monitors that are establishing operational familiarity and a basis for instrument 

performance characterization. In many cases, acid sensitive areas will be strongly influenced 

by regional transport of pollutants which typically is associated with relatively homogeneous 

spatial concentration patterns which allows for a correspondingly greater range of spatial 

representativeness of monitoring sites. Consequently, the expected burden on monitoring 

resources may be realistically dampened by the available infrastructure and expected 

homogeneity of air concentration patterns. A more thorough assessment of the adequacy of 

existing networks is predicated on identification of the area wide boundaries of the acid sensitive 

areas of concern which will initially e developed in the second PAD. 

4.3.3 How do we characterize deposition through monitoring and models? 

Routinely available directly measured precipitation to quantify wet deposition of sulfur 

and nitrogen species is provided through the NADP. Dry deposition is not a directly measured 

variable in routine monitoring efforts. It is important to pursue the development of direct dry 

deposition measurements to improve model parameterizations of deposition processes and 

possibly evolve into routine operations. Estimates of dry deposition based on observations are 

provided through the CASTNET program. However, dry deposition is a calculated value 

represented as the product of ambient concentration (either observed or estimated through air 

quality modeling) and deposition velocity, Amb

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Dry

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Dry

i  CvDep

Deposition velocity is modeled as a mass transfer process through resistance layers 

associated with the canopy, uptake by vegetation, water and soil which collectively are 

influenced by micrometeorology, land surface and vegetation types and species specific 

solubility and reactivity. Dry deposition is calculated through deposition velocity models 

capturing these features and using species specific ambient air concentrations. This approach 

conceptually is similar using either observed or modeled air concentrations. Dry deposition 

estimates from the Community Multi-scale Air Quality (CMAQ) model (EPA, 1999) have been 

used in this assessment to provide spatially more resolved and extensive estimates of dry 

deposition for sulfur and all reactive nitrogen (oxidized and reduced) species (CASTNET does 

not capture important gases such as nitrogen dioxide, ammonia and peroxyacetyl nitrate). All 

of the relevant meteorological, land use, vegetation and elevation data required to estimate 

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deposition velocities are generated or accessible in the CMAQ and/or meteorological preprocessors. 

4.3.2.1 Why are we using CMAQ to model deposition? How are we using it? Why is 

CMAQ the right model to use? What is the spatial and temporal resolution of 

CMAQ? What are the model years? What are the limitations to CMAQ? 

CMAQ provides a platform that allows for a consistent mass accounting approach across 

ambient concentrations and dry and wet deposition values. Recognizing the limitations of 

ambient air networks, CMAQ was used to estimate dry deposition to complement NADP wet 

deposition for MAGIC modeling and for the first-order acidity balance (FAB) critical load 

modeling. CMAQ promotes analytical consistency and efficiency across analyses of multiple 

pollutants. EPA’s Office of Research and Development continues to enhance the underlying 

deposition science in CMAQ. For the purposes of this policy assessment, CMAQ provides a 

consistent platform incorporating the atmospheric and deposition species of interest over the 

entire United States. The caveats and limitations of the use of model predictions are largely 

associated with the general reliance on calculated values, rather than on measurements. Model 

evaluation addressing the comparison of predictions with observed values is addressed in the 

REA and summarized in Chapter 7 of this PA. Currently, there are efforts to improve a number 

of nitrogen related processes in CMAQ, recognizing comparatively less uncertainty with the 

treatment of sulfur. Active areas of model process improvement are in the treatment of 

lightning generated NOX and the transference of nitrogen between atmospheric and terrestrial 

and aquatic media, often referred to as bi-directional flux. Lightning NOx potentially provides 

a significant contribution to wet deposition as the resulting NOX is rapidly entrained into aqueous 

cloud processes. Both the thermodynamics of soil processes and mass transfer of nitrogen 

species across the surface-atmosphere interface is governed by an assortment of temperature, 

moisture, advection and concentration patterns. These processes and mass transfer relationships 

are coupled within the emissions, meteorological, and chemical simulation processes and 

associated surface/vegetation and terrain information incorporated in or accessed by the CMAQ. 

In addition to research activities to improve the characterization of nitrogen-related processes in 

CMAQ, efforts are also underway to improve the general characterization of ammonia emissions 

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which remains as an area of large uncertainty due to limited source data and the ubiquitous 

nature of these emissions. Another challenge for regional/national air quality modeling is 

properly representing the effects on pollutant concentrations, precipitation and therefore 

deposition of variable terrain features, particularly steep mountain-valley gradients and the 

interfaces to wide open basins encountered in the Western United States. 

The CMAQ was used in this assessment because it is the state of science model for 

simulating sources, formation, and fate of nitrogen and sulfur species. In addition to undergoing 

periodic independent scientific peer review, CMAQ bridges the scientific and regulatory 

communities as it is used extensively by EPA for regulatory air quality assessments and rules. 

CMAQ provides hourly estimates of the important precursor, intermediate and secondarily 

formed species associated with atmospheric chemistry and deposition processes influencing 

ozone, particulate matter concentrations and sulfur and nitrogen deposition. Simulations based 

on horizontal spatial scale resolutions of 12 km and 36 km were used in this policy assessment 

for 2002 – 2005. 

4.4 What Is Our Best Characterization of Atmospheric Concentrations Of NOY and 

SOX, and Deposition Of N And S? 

Air quality models and blending of model results and observations are used to 

characterize current environmental state conditions due to the relative sparseness of monitoring 

coverage in sensitive ecosystems as well as gaps in coverage for specific atmospheric species of 

N and S most relevant to deposition, such as NOY, in available monitoring platforms. 

4.4.1 What are the current atmospheric concentrations of reactive nitrogen, NOY, 

reduced nitrogen, NHX, sulfur dioxide, SO2, and sulfate, SO4? 

To provide information for use in characterizing the adequacy of the current standards, we assess 

the best available data for estimating the ambient concentrations of the major sources of 

atmospheric nitrogen and sulfur across the U.S. Acidification and nutrient enrichment processes 

are largely dependent on the cycling of total nitrogen and sulfur species. From an atmospheric 

perspective, it is convenient and consistent with current measurement and modeling frameworks 

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to consider the reduced and oxidized forms of atmospheric nitrogen. Virtually all atmospheric 

sulfur is considered oxidized sulfur in the forms of particulate bound sulfate and gaseous sulfur 

dioxide. In order to assess current concentrations of reactive nitrogen and sulfur, we evaluated 

data available from the existing monitoring networks as well as from the CMAQ model. 

Regarding the monitoring data, there are a number of important issues in understanding the 

measurements of NOY provided by different monitoring networks. In principle, measured NOY is 

based on catalytic conversion of all oxidized species to NO followed by chemiluminescence NO 

detection. We recognize the caveats associated with instrument conversion efficiency and 

possible inlet losses. The CMAQ treats the dominant NOY species as explicit species while the 

minor contributing non-PAN organic nitrogen compounds are aggregated. Atmospheric 

nitrogen and sulfur are largely viewed as regional air quality issues due to the importance of 

chemical conversion of primary emissions into secondarily formed species, a combination of 

ubiquitous sources, particularly mobile source emissions of NOX, and elevated emissions of NOX

and SO2 that aid pollutant mass dispersal and broader physical transport over large distances. In 

effect, the regional nature is due to both transport processes as well as the relatively ubiquitous 

nature of sources combined with chemical processes that tend to form more stable species with 

extended atmospheric lifetimes. This regionalized effect, particularly throughout the eastern 

United States, dominates the overall patterns discussed below of secondarily formed species such 

as sulfate or NOY, which is an aggregate of species with the more aged air masses consisting 

largely of chemically processed air dominated by secondarily formed peroxyacetyl nitrate 

(PAN), particulate nitrate and nitric acid. 

Nationwide maps of CMAQ-predicted 2005 annual average NOY, NHX (NH3 and NH4), 

NH3, NH4, SOX, SO4, and SO2 are provided in Figures 4-5 through 4-11 respectively. Given the 

considerable gaps in air quality observation networks as discussed in the REA and ISA (2008), 

modeled concentration patterns are used here to illustrate national representations of current air 

quality conditions for nitrogen and sulfur. The 2005 model year reflects the most recent 

available simulation for inclusion in this policy assessment. In addition, Figures 4-12 and 4-13 

provide maps of 2005 annual average SO2 and SO4, respectively based on CASTNET 

observations. Site specific annual average 2005 NOY measured concentrations at SLAMS 

(Figure 4-14) are typically are less than 40 ppb. The spatial patterns for the 2005 modeled and 

observed NOY and SO2 concentrations are similar to the 2002 CMAQ-based maps provided in 

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the REA, largely capturing the influence of major source regions throughout the nation. A 

spreading of the oxidized sulfur fields (Figures 4-5 and 4-6), relative to SO2, is consistent with 

sulfate transformation and associated air mass aging and transport. Ammonia and ammonium 

concentration patterns (Figure 4-4) are influenced strongly by the ammonia emissions 

distribution, with marginal spreading associated with the formation of NH4. The NHX fields are 

more strongly influenced by source location, relative to sulfur, based on the fast removal of 

atmospheric ammonia through deposition. However, recent incorporation of ammonia bidirectional flux treatment (see Chapter 7) does reduce NH3 spatial gradients. Total deposition 

for nitrogen and sulfur (Figures 4-15 and 4-16) basically follow the patterns of ambient air 

concentrations. The contribution of reduced nitrogen to total nitrogen deposition (Figure 4-17) 

illustrates the strong influence of agricultural based ammonia emissions, particularly in upper 

midwest and eastern North Carolina. 

The 2005 ambient conditions indicate that the current SO2 and NO2 secondary standards 

are not exceeded (Figures 4-18 and 4-19) in locations where ecological effects have been 

observed, and where critical loads of nitrogen and sulfur are exceeded. This information is 

consistent with the fact that NO2 accounts for only a fraction of NOy, and thus decreases in NO 

and NO2 emissions that result in attaining the current secondary NO2 standard would not be 

expected to fully address deposition of NOY in acid sensitive areas that generally are not 

represented by urban oriented NO2 monitoring locations. The map in Figure 4-20 further 

illustrates this point by showing that the contribution of NO2 to NOY is often less than 50% in 

rural areas. Neither NOY nor NO2 concentrations correlate well with total oxidized nitrogen 

deposition (Figure 4-21), based on the annual average values in each 12 km CMAQ grid cell. 

The lack of correlation between NOY and nitrogen deposition is largely due to the inclusion of 

NOY species with low deposition velocity, primarily NO2. While NO2 does not reflect the 

majority of oxidized nitrogen in rural environments, the species remains a significant contributor 

to the total ambient NOY budget. In contrast, species with high deposition velocities such as 

nitric acid correlate well with oxidized nitrogen deposition. The temporal correlations between 

NOY and deposition are likely to be similar to the nitric acid to deposition relationship when 

averaged over larger a spatial area as the influence of species with low deposition velocities is 

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Figure 4-5. 2005 CMAQ modeled annual average NOY (ppb; see Table 1-1 for unit 

conversions). 

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Figure 4-6. 2005 CMAQ modeled annual average total reduced nitrogen (NHX) (as ug/m3

nitrogen – see Table 1-1 for unit conversions) 

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 Figure 4-7. 2005 CMAQ modeled annual average ammonia, NH3, (as ug/m3 N; see Table 

1-1 for unit conversions ) 

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Figure 4-8. 2005 CMAQ modeled annual average ammonium, NH4, (as ug/m3 N; see 

Table 1-1 for unit conversions) 

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Figure 4-9. 2005 CMAQ modeled annual average SOX, (as ug/m3

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see Table 1-1 for unit conversions). 

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Figure 4-10. 2005 CMAQ modeled annual average SO2 (as ug/m3

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Figure 4-11. 2005 CMAQ modeled annual average SO4 (as ug/m3

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Figure 4-12. 2005 annual average sulfur dioxide concentrations based on CASTNET 

generated by the Visibility Information Exchange Web Sysytem (VIEWS). 

See Table 1-1 for unit conversions

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Figure 4-13 2005 annual average sulfate concentrations based on CASTNET generated 

by the Visibility Information Exchange Web Sysytem (VIEWS). [interpolating 

relative sparse data can produce unrealistic concentration plumes as demonstrated 

in the central U.S.] see Table 1-1 for unit conversions. 

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Figure 4-14. Annual average 2005 NOY concentrations from reporting stations in the Air 

Quality System (AQS). see Table 1-1 for unit conversions. 

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Figure 4-15. 2005 CMAQ modeled oxidized nitrogen deposition (kgN/ha-yr). see Table 1-

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Figure 4-16. 2005 CMAQ modeled oxidized sulfur deposition (kgS/ha-yr). see Table 1-1 

for unit conversions

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Figure 4-17. 2005 CMAQ derived annual average ratio of reduced to total nitrogen 

deposition. 

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Figure 4-18. Three hour average maximum 2005 SO2 concentrations based on the SLAMS 

reporting to EPA’s Air Quality System (AQS) data base. The current SO2

secondary standard based on the maximum 3 hour average value is 500 ppb, a 

value not exceeded. While there are obvious spatial gaps, the majority of these 

stations are located to capture maximum values generally in proximity to major 

sources and high populations. Lower relative values are expected in more remote 

acid sensitive areas. see Table 1-1 for unit conversions

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Figure 4-19 Annual average 2005 NO2 concentrations based on the SLAMS reporting to 

EPA’s Air Quality System (AQS) data base. The current NO2 secondary 

standard is 53 ppb, a value well above those observed. While there are obvious 

spatial gaps, the stations are located in areas of relatively high concentrations in 

highly populated areas. Lower relative values are expected in more remote acid 

sensitive areas. see Table 1-1 for unit conversions 

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Figure 4-20 2005 CMAQ derived annual average ratio of (NOY – NO2)/NOY. The fraction 

of NO2 contributing to total NOY generally is less than 50% in the Adirondack 

and Shenandoah case study areas. The ratio (dimensionless – scale units of ppbv 

are an automated output) reflects the relative air mass aging associated with 

transformation of oxidized nitrogen beyond NO and NO2 as one moves from 

urban to rural locations. 

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Figure 4-21 Scatter plots of total oxidized nitrogen deposition with average annual NO2, 

HNO3 and NOY concentrations on each 12 km2 grid based on 2005 CMAQ 

results for the Adirondack region. 

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4.5 Are Adverse Effects On The Public Welfare Occurring Under Current Air Quality 

Conditions For NO2 And SO2 And Would They Occur If The Nation Met The 

Current Secondary Standards? 

In the previous sections we have established that almost all areas of the U.S. were at 

concentrations of SO2 and NO2 below the levels of the current standards. In many locations, SO2

and NO2 concentrations are substantially below the levels of the standards. This pattern suggests 

that levels of deposition and any effects on ecosystems due to deposition of NOx and SOx under 

recent conditions are occurring even though areas meet or are below current standards. In this 

section we focus on summarizing the evidence of effects occurring at deposition levels consistent 

with recent conditions. 

The ISA summarizes the available studies of relative nitrogen contribution and finds that 

in much of the U.S., NOX contributes from 50 to 75 percent of total atmospheric deposition 

relative to total reactive nitrogen that includes oxidized and reduced nitrogen species.[ISA 

Section 2.8.4]. Although the proportion of total nitrogen loadings associated with atmospheric 

deposition of nitrogen varies across locations (N deposition in the eastern U.S. includes locations 

with greater than 9 kg N/ha-yr, and in the central U.S. high deposition locations with values on 

the order of 6 to 7 kg N/ha-yr), the ISA indicates that atmospheric N deposition is the main 

source of new anthropogenic N to most headwater streams, high elevation lakes, and low-order 

streams. Atmospheric N deposition contributes to the total N load in terrestrial, wetland, 

freshwater, and estuarine ecosystems that receive N 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 (U.S. EPA, 2000). 

Atmospheric concentrations of SOx account for nearly all S deposition in the US. For the 

period 2004–2006, mean S 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 S deposition in the U.S. west of the 100th meridian was 

relatively low, with all recording stations reporting <2 kg S/ha-yr and many reporting <1 kg 

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S/ha-yr. S was primarily deposited in the form of wet SO4

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New scientific evidence exists to address each of the areas of uncertainty raised in the 

previous reviews (summarized above in section 1.4). Based on the new evidence, the current 

ISA concludes that: 

(1) The evidence is sufficient to infer a causal relationship between acidifying 

deposition (to which both NOX and SOX contribute) and effects on 

biogeochemistry related to terrestrial and aquatic ecosystems; and biota in 

terrestrial and aquatic ecosystems. 

(2) The evidence is sufficient to infer a causal relationship between N deposition, 

to which NOX and NHX contribute, and the alteration of A) biogeochemical 

cycling of N and carbon in terrestrial, wetland, freshwater aquatic, and coastal 

marine ecosystems; B) biogenic flux of methane (CH4), and N2O in terrestrial 

and wetland ecosystems; and C) species richness, species composition, and 

biodiversity in terrestrial, wetland, freshwater aquatic and coastal marine 

ecosystems. 

(3) The evidence is sufficient to infer a causal relationship between S deposition 

and increased Hg methylation in wetlands and aquatic environments. 

Subsequent to the previous review of the NOX secondary standard, a great deal of 

information on the contribution of atmospheric deposition associated with ambient NOX has 

become available. In Chapter 3 of the REA a thorough assessment is provided of the 

contribution of NOX to nitrogen deposition throughout the U.S., and the relative contributions of 

ambient NOX and reduced forms of nitrogen. Staff concludes that based on that analysis, 

ambient NOX is a significant component of atmospheric nitrogen deposition, even in areas with 

relatively high rates of deposition of reduced nitrogen. In addition, staff concludes that 

atmospheric deposition of oxidized nitrogen contributes significantly to total nitrogen loadings in 

nitrogen sensitive ecosystems. 

As discussed throughout the REA document, there are several key areas of risk that are 

associated with ambient concentrations of NOX and SOX. As noted earlier, in previous reviews 

of the NOX and SOX secondary standards, the standards were designed to protect against direct 

exposure of plants to ambient concentrations of the pollutants. A significant shift in 

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understanding of the effects of NOX and SOX 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 risks of adverse effects to public welfare are those 

related to deposition of NOX and SOX to both terrestrial and aquatic ecosystems. These risks fall 

into two categories: acidification and nutrient enrichment. These made up the emphasis of the 

REA, and are most relevant to evaluating the adequacy of the existing standards in protecting 

public welfare from adverse ecological effects. 

4.5.1 To what extent do the current NOX and SOX secondary standards provide 

protection from adverse effects associated with deposition of atmospheric NOX and SOX

which results in acidification in sensitive aquatic and terrestrial ecosystems? 

The focus of the REA case studies was on determining whether deposition of sulfur and 

oxidized nitrogen in locations where ambient NOX and SOX was at or below the current 

standards was resulting in acidification and related effects. This review has focused on 

identifying ecological indicators that can link atmospheric deposition to ecological effects 

associated with acidification. NOX and SOX contribute to acidification in both aquatic and 

terrestrial ecosystems, although the indicators of effects differ. Although there are some 

geographic areas with both terrestrial and aquatic ecosystems that are vulnerable to acidification, 

the case study areas do not fully overlap. The locations of the case studies evaluated in the REA 

are shown on Figure 4-20. 

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Figure 4-22 National map highlighting the nine case study areas evaluated in the REA. 

4.5.1.1 Aquatic Acidification 

Based on the case studies conducted for lakes in the Adirondacks and streams in 

Shenandoah National Park, staff concludes that there is significant risk to acid sensitive aquatic 

ecosystems at atmospheric concentrations of NOX and SOX at or below the current standards. 

This conclusion is based on application of the MAGIC model to estimate the effects of 

deposition at levels consistent with atmospheric NOX and SOX concentrations that are at or 

below the current standards. An important ecological indicator for aquatic acidification effects is 

acid neutralizing capacity (ANC) of a waterbody, and the case study focused on evaluating 

whether locations were likely to be below critical values of ANC given deposition levels 

associated with NOx and SOx atmospheric concentrations that meet the current standards. In 

addition, the case studies assessed the ecological effects and some of the known ecosystem 

services that are associated with different levels of ANC in order to associate levels of ANC with 

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measures of public welfare that may be adversely affected by deposition levels consistent with 

atmospheric concentrations of NOX and SOX that meet the current standards. 

 Staff concludes that the evidence and risk assessment support strongly a relationship 

between atmospheric deposition of NOX and SOX and loss of ANC in sensitive ecosystems, and 

that ANC is an excellent indicator of aquatic acidification. Staff also concludes that at levels of 

deposition associated with NOX and SOX 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 (REA Section 4). 

Many locations in sensitive areas of the U.S. have ANC levels below benchmark levels 

for ANC classified as severe, elevated, or moderate concern (see 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, and 28 percent of lakes had deposition levels exceeding the critical load for an 

ANC of 20 μeq/L (REA 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 levels, 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 (REA Figure 4.2-1). At levels 

below 20 μeq/L, populations of sensitive species, such as brook trout, may decline significantly 

during episodic acidification events. When extrapolated to the full population of lakes in the 

Adirondacks area using weights based on the EMAP probability survey (REA 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 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 significant 

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numbers of sensitive streams in the Shenandoah area are at risk of adverse impacts on fish 

populations under recent conditions. Many other streams in the Shenandoah area are 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 (REA 4.2.6.2). 

In the ISA it is noted that significant portions of the U.S. are acid sensitive, and that 

current deposition levels exceed those that would allow recovery of the most acid sensitive lakes 

in the Adirondacks (ISA ES). In addition, because of past loadings, areas of the Shenandoah are 

sensitive to current deposition levels (ISA ES). 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. 

While most (99 percent) of stream kilometers in the U.S. are not chronically acidified 

under current conditions, a recent survey found sensitive streams in many locations in the U.S., 

including the Appalachian mountains, the Coastal Plain, and the Mountainous West (ISA 

Section 4.2.2.3). In these sensitive areas, between 1 and 6 percent of stream kilometers are 

chronically acidified. 

The ISA notes that “consideration of episodic acidification greatly increases the extent 

and degree of estimated effects for acidifying deposition on surface waters.” (ISA 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 (ISA 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. (ISA Section 

3.2.1.6). 

It can therefore be concluded that recent levels of NOX and SOX are associated with 

deposition that leads to ANC values below benchmark values known to cause ecological harm in 

sensitive aquatic systems, including lakes and streams in multiple areas of the U.S. These 

changes are known to have impacts on ecosystem services including recreational fishing which is 

discussed along with other services in Chapter 3. While other ecosystem services (e.g. habitat 

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provisioning, subsistence fishing, and biological control as well as many others) are potentially 

affected by reductions in ANC, confidence in the specific translation of ANC values to these 

additional ecosystem services is much lower. 

4.5.1.2 Terrestrial Acidification 

 Based on the case studies on sugar maple and red spruce habitat, staff concludes that 

there is significant risk to sensitive terrestrial ecosystems from acidification at atmospheric 

concentrations of NOX and SOX at or below the current standards. This conclusion is based on 

application of the simple mass balance model to deposition levels associated with NOX and SOX

concentrations at or below the current standards. The ecological indicator selected for terrestrial 

acidification is the base cation to aluminum ratio (BC:Al), which has been linked to tree health 

and growth. The results of the REA strongly support a relationship between atmospheric 

deposition of NOX and SOX and BC:Al, and that BC:Al is a good indicator of terrestrial 

acidification. At levels of deposition associated with NOX and SOX 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.) and many 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 (REA 

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 

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of 0.6. However, there was large variability across states. In New Jersey, 67 percent of plots 

exceeded the critical load for a BC:Al ratio of 1.2, while in several states on the outskirts of the 

range for sugar maple (e.g. Arkansas, Illinois) no plots exceeded the critical load for a BC:Al 

ratio of 1.2. For red spruce, overall 5 percent of plots exceeded the critical load for a BC:Al ratio 

of 1.2, and 3 percent exceeded the critical load for a BC:Al ratio of 0.6. In the major red spruce 

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. 

The ISA reported one study (McNulty, 1997) that estimated 15 percent of U.S. forest 

ecosystems exceeded the critical loads for acidity for N and S deposition by >250 eq/ha/year 

under current conditions (ISA Section 4.2.1.3) . Staff concludes that this represents a significant 

portion of sensitive terrestrial ecosystems. 

It can therefore be concluded that recent levels of NOX and SOX are associated with 

deposition that leads to BC:Al values below benchmark values that cause ecological harm in 

some sensitive terrestrial ecosystems. While effects are more widespread for sugar maple, there 

are locations with low to intermediate levels of protection from effects on both sugar maple and 

red spruce. While there are many other ecosystem services, including timber production, natural 

habitat provision, and regulation of water, climate, and erosion, potentially affected by 

reductions in BC:Al, linkages of BC:Al values to these additional ecosystem services is on the 

whole not well understood. 

4.5.2 To what extent does the current NOX secondary standard provide protection from 

adverse effects associated with deposition of atmospheric NOX, which results in nutrient 

enrichment effects in sensitive aquatic and terrestrial ecosystems? 

Nutrient enrichment effects are due to nitrogen loadings from both atmospheric and nonatmospheric sources. Evaluation of nutrient enrichment effects requires an understanding that 

nutrient inputs are essential to ecosystem health. The 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. In certain limited situations, additions of 

nitrogen can increase rates of growth, and these increases can have short term benefits in certain 

managed ecosystems. As noted earlier, this review of the standards is focused on unmanaged 

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ecosystems. As a result, in assessing adequacy of the current standards, we are focusing on the 

adverse effects of nutrient enrichment in unmanaged ecosystems. However, the following 

discussion provides a brief assessment of effects in managed ecosystems. 

Impacts of nutrient enrichment in managed ecosystems may be positive or negative 

depending on the levels of nutrients from other sources in those areas. Positive effects can occur 

when crops or commercial forests are not receiving enough nitrogen nutrients. Nutrients 

deposited on crops from atmospheric sources are often referred to as passive fertilization. 

Nitrogen is a fundamental nutrient for primary production in both managed and unmanaged 

ecosystems. Most productive agricultural systems require external sources of nitrogen in order 

to satisfy nutrient requirements. Nitrogen uptake by crops varies, but typical requirements for 

wheat and corn are approximately 150 kg/ha-yr and 300 kg/ha-yr, respectively (NAPAP, 1990). 

These rates compare to estimated rates of passive nitrogen fertilization in the range of 0 to 5.5 

kg/ha-yr (NAPAP, 1991). 

Information on the effects of changes in passive 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 (ISA 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 NOx deposition will be variable depending on the 

forest management practices employed in a given forest ecosystem (ISA Annex C C.6.3) 

Increases in the availability of nitrogen in N-limited forests via atmospheric deposition could 

increase forest production over large non-managed areas, but the evidence is mixed, with some 

studies showing increased production and other showing little effect on wood production (ISA 

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 (ISA 3.3.9).

In managed agricultural ecosystems, nitrogen inputs from atmospheric NOx comprise a 

small fraction (less than 3 percent) of total nitrogen inputs, which include commercially applied 

fertilizers as well as applications of composted manure. And because of the temporal and spatial 

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variability in atmospheric deposition of NOx, it is unlikely that farmers would alter their 

fertilization decisions based on expected nitrogen inputs from NOX. And, in some locations, 

farmers need less nitrogen inputs due to production of excess nitrogen through livestock. In 

some locations, nitrogen production through livestock waste exceeds the absorptive capacity of 

the surrounding land, and as such, excess nitrogen from deposition of NOX in those locations 

reduces the capacity of the system to dispose of excess nitrogen, potentially increasing the costs 

of waste management from livestock operations (Letson and Gollehon, 1996). A USDA 

Economic Research Service report found that in 1997, 68 counties with high levels of confined 

livestock production had manure nitrogen levels that exceed the assimilative capacity of the 

entire county’s crop and pasture land (Gollehon et al, 2001). In those locations, additional 

nitrogen inputs from NOX deposition will result in excess nitrogen, leading to nitrogen leaching 

and associated effects that adversely effect ecosystems. 

4.5.3 Aquatic Nutrient Enrichment 

The REA case studies focused on coastal estuaries and revealed that while current 

ambient loadings of atmospheric NOX 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 as the ecological 

indicator for aquatic nutrient enrichment an index of eutrophication known as the Assessment of 

Estuarine Trophic Status Eutrophication Index (ASSETS EI). This index is a six level index 

characterizing overall eutrophication risk in a waterbody. This indictor is not sensitive to 

relatively large changes in nitrogen deposition. In addition, this type of indicator does not reflect 

the impact of nitrogen deposition in conjunction with other sources of nitrogen. 

For example, if NOx deposition is contributing nine tenths of the nitrogen loading 

required to move a waterbody from an ASSETS EI category of “moderate” to a category of 

“poor”, zeroing out NOX deposition will have no impact on the ASSETS EI value. However, if 

an area were to decide to put in place decreases in nitrogen loadings to move that waterbody 

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from “poor” to “moderate,” the area would have to reduce the full amount of the loadings 

through other sources if atmospheric deposition were not considered. Thus, the adverse impact 

of atmospheric nitrogen is in its contribution to the overall loading, and reductions in NOX will 

decrease the amount of reductions from other sources of nitrogen loadings that would be required 

to move from a lower ASSETS EI category to a higher category. NOX deposition can also be 

characterized as reducing the risk of a waterbody moving from a higher ASSETS EI category to 

a lower category, by reducing the vulnerability of that waterbody to increased loadings from 

non-atmospheric sources. 

Based on the above considerations, staff preliminarily concludes that the ASSETS EI is 

not an appropriate ecological indicator for estuarine aquatic eutrophication. Staff further 

concludes 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 NOX. As a result, staff is unable to make a determination as to the adequacy of the 

existing secondary NOX standard in protecting public welfare from N 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/hayr which are well within current ambient levels are known to cause changes in species 

composition in diatom communities indicating impaired water quality (ISA Section 3.3.5.3). It 

then seems apparent then that the existing secondary standard for NOX does not protect such 

ecosystems and their resulting services from impairment. 

4.5.4 Terrestrial Nutrient Enrichment 

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. These benchmarks are 

discussed more thoroughly in Chapter 5 of the REA. Large areas of the country appear to be 

experiencing deposition above these benchmarks for example, Fenn et al. (2008) 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 

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Colorado Front Range of the Rocky Mountain National Park support these findings and are 

summarized in Chapter 6.0 of the Risk and Exposure Assessment. These three values (3.1, 5.2, 

and 10.2 kg/ha-yr) are one set of ecologically meaningful benchmarks for the mixed conifer 

forest (MCF) of the pacific coast regions. Nearly all of the known sensitive communities receive 

total nitrogen deposition levels above the 3.1 N kg/ha-yr ecological benchmark according to 

the12 km, 2002 CMAQ/NADP data, with the exception of the easternmost Sierra Nevadas. 

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 (CSS) are also known to be sensitive to community 

shifts caused by excess nitrogen loadings. Wood et al. (2006) 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 (Wood et al., 2006). It is assumed that 3.3 kg 

N/ha-yr is near the point where nitrogen is no longer limiting in the CSS community. 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, staff 

concludes that recent conditions where NOX ambient concentrations are at or below the current 

NOX secondary standards are not adequate to protect against anticipated adverse impacts from N 

nutrient enrichment in sensitive ecosystems. 

4.6 To What Extent Do The Current NOX And/Or SOX Secondary Standards Provide 

Protection From Other Ecological Effects (Eg. Mercury Methylation) Associated 

With The Deposition Of Atmospheric NOX, And/Or SOX? 

It is stated in the ISA (ISA Sections 3.4.1 and 4.5) that mercury is a highly neurotoxic 

contaminant that enters the food web as a methylated compound, methylmercury. 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 methylmercury can be produced in the absence of sulfate. Once methylmercury is 

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

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fish can only occur where substantial amounts of methylmercury are present. Current evidence 

indicates that in watersheds where mercury is present, increased SOX deposition very likely 

results in additional production of methylmercury which leads to greater accumulation of MeHg 

concentrations in fish (Munthe et al, 2007; Drevnick et al., 2007). 

The production of meaningful amounts of methylmercury (MeHg) requires the presence 

of SO4

2- and mercury, and where mercury is present, increased availability of SO42- 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 

methylmercury requires the presence of sulfate and mercury, but the amount of methylmercury 

produced varies with oxygen content, temperature, pH, and supply of labile organic carbon (ISA 

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 (ISA Section 3.4). Watersheds with conditions known to be conducive to mercury 

methylation can be found in the northeastern United States and southeastern Canada. The 

relationship between sulfur and methylmercury production is addressed qualitatively in Chapter 

6 of the Risk and Exposure Assessment. 

With respect to sulfur deposition and mercury methylation, the final ISA determined: The 

evidence is sufficient to infer a causal relationship between sulfur deposition and increased 

mercury methylation in wetlands and aquatic environments. However, staff did not conduct a 

quantitative assessment of the risks associated with increased mercury methylation under current 

conditions. As such, staff are unable to make a determination as to the adequacy of the existing 

SO2 standards in protecting against welfare effects associated with increased mercury 

methylation. 

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4.7 REFERENCES: 

B.J. Finlayson-Pitts and J.N. Pitts, 2000, Chemistry of the Upper and Lower Troposhere, 

Academic Press, San Diego, CA 

Drevnick, P.E., D.E. Canfield, P.R. Gorski, A.L.C. Shinneman, D.R. Engstrom, D.C.G. Muir, 

G.R. Smith, P.J. Garrison, L.B. Cleckner, J.P. Hurley, R.B. Noble, R.R. Otter, and J.T. 

Oris. 2007. Deposition and cycling of sulfur controls mercury accumulation in Isle 

Royale fish. Environmental Science and Technology 41(21):7266–7272. 

Fenn, M.E., S. Jovan, F. Yuan, L. Geiser, T. Meixner, and B.S. Gimeno. 2008. Empirical and 

simulated critical loads for nitrogen deposition in California mixed conifer forests. 

Environmental Pollution 155(3):492–511. 

Munthe, J., R.A. Bodaly, B.A. Branfireun, C.T. Driscoll, C.C. Gilmour, R. Harris, M. Horvat, M. 

Lucotte, and O. Malm. 2007. Recovery of mercury-contaminated fisheries. Ambio 36:33–

44. 

Scheffe, R.D., P. A. Solomon, R. Husar, T. Hanley, M. Schmidt, M. Koerber , M. Gilroy, J. 

Hemby, N. Watkins, M. Papp, J. Rice, J. Tikvart andR. Valentinetti, The National 

Ambient Air Monitoring Strategy: Rethinking the Role of National Networks, JAWMA, 

ISSN:1047-3289 J. Air & Waste Manage. Assoc. 2009, 59:579–590 DOI:10.3155/1047-

3289.59.5.579 

U.S. EPA (Environmental Protection Agency). 1982. Review of the National Ambient Air Quality 

Standards for Sulfur Oxides: Assessment of Scientific and Technical Information. 

OAQPS Staff Paper. EPA-450/5-82-007. U.S. Environmental Protection Agency, Office 

of Air Quality Planning and Standards, Research Triangle Park, NC. 

U.S. EPA (Environmental Protection Agency). 1995. Review of the National Ambient Air Quality 

Standards for Nitrogen Dioxide: Assessment of Scientific and Technical Information. 

OAQPS Staff Paper. EPA-452/R-95-005. U.S. Environmental Protection Agency, Office 

of Air Quality Planning and Standards, Research Triangle Park, NC. September. 

U.S. EPA. (Environmental Protection Agency). 2000. Deposition of Air Pollutants to the Great 

Waters: Third report to Congress. (EPA-453/R-00-005) 

http://www.epa.gov/air/oaqps/gr8water/3rdrpt/index.html; Research Triangle Park, NC; 

Office of Air Quality Planning and Standards; 

 U.S. Environmental Protection Agency 

U.S. EPA (Environmental Protection Agency). 2008. Integrated Science Assessment (ISA) for 

Oxides of Nitrogen and Sulfur–Ecological Criteria (Final Report). EPA/600/R08/082F. U.S. Environmental Protection Agency, National Center for Environmental 

Assessment–RTP Division, Office of Research and Development, Research Triangle 

Park, NC. Available at http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=201485. 

Wood, Y., T. Meixner, P.J. Shouse, and E.B. Allen. 2006. Altered Ecohydrologic response 

drives native shrub loss under conditions of elevated N-deposition. Journal of 

Environmental Quality 35:76–92. 





























































































6. CO-PROTECTION FOR OTHER EFFECTS USING STANDARDS TO 

PROTECT AGAINST ACIDIFICATION

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Discussion in this Policy Assessment on the NOX and SOX secondary standard has, to this 

point, centered on the level of protection of aquatic ecosystems against acidification from 

atmospheric deposition of NOX and SOX. This chapter focuses on the co-protection such a 

standard could achieve for other ecological effects, including terrestrial acidification, terrestrial 

nutrient enrichment, and estuarine eutrophication. 

6.1 To What Extent Would A Standard Specifically Defined To Protect Against Aquatic 

Acidification Likely Provide Protection From Terrestrial Acidification?

To understand the level of co-protection a NOX and SOX secondary standard based on 

aquatic acidification can provide for terrestrial ecosystems, EPA staff conducted an analysis to 

compare the critical acid loads for aquatic and terrestrial components of watersheds in the eastern 

United States. Aquatic critical acid loads are an integrated function of the chemistry of runoff 

from stream and lake waters, and the biogeochemical processes that occur within the aquatic and 

terrestrial components of the entire watershed. Terrestrial critical acid loads, however, are 

largely determined by the conditions and processes that occur in the root zone of the soil profile 

of the terrestrial systems of a watershed. Therefore, it is possible to have different critical acid 

load values for aquatic and terrestrial ecosystems within the same watershed. 

For the comparative analysis of aquatic and terrestrial critical acid loads, aquatic critical 

acid loads were selected to protect for an acid neutralizing capacity (ANC) of 50μeq/L, and were 

taken directly from the Risk and Exposure Assessment (REA) (2009). The 50μeq/L ANC value 

was one of three values modeled in the REA (2009) for aquatic acidification. The terrestrial 

critical acid loads in this comparative analysis were selected to protect for either a terrestrial base 

cation to aluminum molar ratio (Bc:Al) of 1.2 or 10.0. The Bc:Al ratio of 10.0 is more 

conservative, as it provides greater protection against the impacts of acidification on cation 

availability and aluminum toxicity in the soil solution. The terrestrial critical loads were 

calculated using the Simple Mass Balance (SMB) method outlined in the REA (2009) and input 

values averaged across the area of each watershed. 

September, 2010 6-1 Draft – Do Not Quote or Cite 
Aquatic and terrestrial critical acid loads were compared in 16 watersheds from each of 

the two aquatic acidification Case Study Areas, the Adirondacks and the Shenandoah, identified 

in the REA (2009). For each Case Study Area, four watersheds were randomly selected from 

each of the four aquatic acidification sensitivity classes reported in the REA (2009). Those four 

sensitivity classes are “highly sensitive”, “moderately sensitive”, “low sensitivity”, and “not 

sensitive”. In order for a watershed to be classified as one of these four classes, it had to contain 

at least one lake or stream with that sensitivity class designation. The Adirondacks Case Study 

Area contained watersheds representing all four sensitivity classes, and the 16 watersheds that 

were selected for the analysis contained a total of 29 lakes. However, in the Shenandoah Case 

Study Area, there were a limited number of watersheds in the “low” and “not sensitive” classes. 

Therefore, only one of the 16 randomly selected watersheds contained a “low” and a “not 

sensitive” stream. In total, there were 20 streams located in the 16 Shenandoah watersheds 

selected for the comparative analysis. In each of the 32 watersheds (16 Adirondacks plus 16 

Shenandoah), the terrestrial critical acid loads were calculated as a single value for the entire 

watershed. These terrestrial critical acid loads were then compared to the aquatic critical acid 

loads for the lakes and streams within each watershed to determine whether the aquatic or 

terrestrial critical acid load provided greater protection against acidifying nitrogen and sulfur 

deposition. Appendix B of this Policy Assessment document provides a full description of the 

methods and results of this comparative analysis. 

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Results of the comparison between the aquatic critical acid load (ANC = 50 μeq/L) and 

the terrestrial critical acid loads (Bc:Al 1.2 and 10.0) for the 32 watersheds are presented in 

Tables 6.1 and 6.2. In the 16 Adirondack watersheds, 13 of the 29 lakes had aquatic critical 

acid loads that were lower (more protective) than the terrestrial critical acid loads when a Bc:Al 

ratio of 10.0 was used. Based on terrestrial critical acid loads determined with a Bc:Al ratio of 

1.2, 21 of the 29 lakes in the Adirondacks had aquatic critical acid loads lower than the terrestrial 

critical acid loads. More importantly, for the terrestrial critical acid loads determined with a 

Bc:Al ratio of 10.0, 13 of the 16 lakes in the Adirondacks classified as “highly” and 

“moderately” sensitive to acidification had aquatic critical acid loads lower than the terrestrial 

critical acid loads, and all 16 lakes in these two sensitivity classes had critical acid loads lower 

than the terrestrial loads determined with a Bc:Al of 1.2 The watersheds within the Shenandoah 

region showed similar results (Table 6.1). 

September, 2010 6-2 Draft – Do Not Quote or Cite 
Table 6-1. Results of the comparison of lake and stream aquatic critical loads (ANC of 50 

μeq/L) to terrestrial critical loads (Bc:Al molar ratios of 10.0 in soil solution) calculated for 

the full watershed in each of the 16 watersheds in the Adirondacks and Shenandoah Case 

Study Areas. The tabular results show the number of times the aquatic acidification critical load 

would provide more protection than the terrestrial acidification critical load. 

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Case Study Area Watershed Sensitivity to Aquatic Acidification 

Highly Sensitive Moderately Sensitive Low Sensitivity Not Sensitive

Adirondacks 7 of 7 6 of 9 0 of 7 0 of 6 

Shenandoh 14 of 14 5 of 5 0 of 1 0 of 1 

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Table 6-2. Results of the comparison of lake and stream aquatic critical loads (ANC of 50 

μeq/L) to terrestrial critical loads (Bc:Al molar ratios of 1.2 in soil solution) calculated for 

the full watershed in each of the 16 watersheds in the Adirondacks and Shenandoah Case 

Study Areas. The tabular results show the number of times the aquatic acidification critical load 

would provide more protection than the terrestrial acidification critical load. 

Case Study Area Watershed Sensitivity to Aquatic Acidification 

Highly Sensitive Moderately Sensitive Low Sensitivity Not Sensitive

Adirondacks 7 of 7 9 of 9 5 of 7 0 of 6 

Shenandoh 14 of 14 5 of 5 0 of 1 0 of 1 

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In summary, terrestrial and aquatic critical acid loads were compared for watersheds in 

the Adirondack and Shenandoah Case Study Areas. Results indicated that, in general, the 

aquatic critical acid loads were lower and therefore offered greater protection to the watershed 

than did the terrestrial critical acid loads. In situations where the terrestrial critical acid loads 

were lower (i.e., more protective) than the aquatic critical acid loads, the lakes or streams in the 

watershed were often rated as having “low sensitivity” or “not sensitive” to acidifying nitrogen 

and sulfur deposition. Conversely, when the waterbodies were more sensitive to deposition 

(“highly sensitive” or “moderately sensitive”), the aquatic critical acid loads generally provided a 

September, 2010 6-3 Draft – Do Not Quote or Cite 
greater level of protection against acidifying nitrogen and sulfur deposition in the watershed. It 

is uncertain whether these results would be consistent for the rest of the country. 

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3 6.2 To What Extent Would A Standard Specifically Defined To Protect Against Aquatic 

Acidification Likely Provide Protection From Terrestrial Nutrient Enrichment? 

The figure below summarizes the terrestrial nutrient enrichment effects from nitrogen 

discussed in the REA (2009). The REA reported these benchmarks as kg/ha/yr. To convert to 

meq/m2/yr each benchmark number must be multiplied by 7.14. The range in meq/m2/yr thus 

becomes 10.7 meq/m2/yr for the low end benchmark of changes to diatom community structure 

to 321 meq/m2/yr at the high end benchmark associated with nitrate leaching, high foliar 

nitrogen, and high NO emissions in southern California. 

For each depositional load that is considered for aquatic acidification, whether it is a 

national number or a sensitivity based number, it can be compared against the chart in figure 6.1 

to understand the level of protection offered in individual parts of the country where these 

studies were conducted. For example referring back to the figure under option 1 in section 

5.3.2, the depositional load selected to encompass 90% of the critical loads on a national basis to 

protect for an ANC of 50 is described as a tradeoff curve between sulfur and nitrogen. If sulfur 

were zero then the maximum nitrogen deposition would be 79 meq/m2/yr or 11 kg/ha/yr. 

Comparing this maximum nitrogen deposition number to the benchmarks in figure 6.1 shows 

that the depositional load would provide some protection against leaching in northeast forests, 

but would have to be lower to protect California coastal sage scrub, lichens in mixed conifer 

forests, alpine lake communities, and Minnesota grasslands. 

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Figure 6-1 Benchmarks of atmospheric nitrogen deposition for several ecosystem 

indicators. 

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2 6.3 To What Extent Would A Standard Specifically Defined To Protect Against Aquatic 

Acidification Likely Provide Protection From Aquatic Nutrient Enrichment? 

The REA (2009) found that deposition of reactive nitrogen contributed to eutrophication 

of estuaries; however, it was also noted that atmospheric deposition of nitrogen is only part of 

the total nitrogen load to the estuaries. Due to the complications of separating out the effects of 

atmospheric deposition from the effects of other nitrogen loads, CASAC did not recommend that 

a secondary NAAQS be set to specifically protect against estuarine eutrophication at this time. 

As described in the REA (2009), the Chesapeake Bay is one national estuary that is 

suffering from eutrophication. In issuing his Executive Order on the Chesapeake Bay, President 

Obama recognized that the Bay watershed is one of our nation’s greatest treasures and must be 

protected and restored. To that end, EPA is proposing a nitrogen total maximum daily load 

(TMDL) for the Chesapeake Bay. The TMDL will contain a specific air allocation for nitrogen 

deposition. The allocations that were provided to the states included assumptions that air 

deposition levels of nitrogen would be reduced to 14.9 million pounds per year to the tidal waters 

and to 323 million pounds to the watershed by the year 2020. According to the Chesapeake Bay 

Program Office, the tidal waters have a surface area of 4,479 square miles and the watershed is 

64,216 square miles. This means that in 2020, the TMDL currently calls for nitrogen deposition 

levels to the combined bay and watershed to be reduced to 337.9 million pounds/68,695 square 

miles/yr, which is equivalent to 8.6 kg/ha/yr or 61 meq/m2/yr. As in Section 6.2, this number 

can be compared to the maximum depositional load of 79 meq/m2/yr. If sulfur were zero on the 

national tradeoff curve from the figure under option 1 of section 5.3.2, then the allowed 

depositional load of 79 meq/m2/yr would not meet the Chesapeake Bay TMDL as currently 

envisioned. 

 

September, 2010 6-6 Draft – Do Not Quote or Cite 
September, 2010 6-7 Draft – Do Not Quote or Cite 

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6.4 REFERENCE 

US EPA (United States Environmental Protection Agency). 2009. Risk and Exposure Assessment 

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

Nitrogen and Sulfur. Final. U.S. Environmental Protection Agency, Office of Research 

and Development, National Center for Environmental Assessment, Research Triangle 

Park, NC. September. 
7 EVALUATION OF UNCERTAINTY AND VARIABILITY IN THE 

CONTEXT OF AN AAPI STANDARD, INCLUDING MODEL 

EVALUATION, SENSITIVITY ANALYSES, AND ASSESSMENT OF 

INFORMATION GAPS 

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7.1 INTRODUCTION AND PURPOSE 

This chapter provides discussions of results of analyses and assessments intended to 

address the relative confidence associated with many of the individual and combined 

components of the linked atmospheric-ecological effects system described in Chapter 5, as well 

as important uncertainties in the scientific evidence that should be considered in developing 

options for the standard. This chapter is intended to integrate a variety of analyses related to the 

sensitivity of the models and model components to uncertainty and variability, and place the 

results of those analyses within the context of the conclusions that can be drawn regarding the 

components of the AAPI. These components include ecosystem effects; dose-response 

relationships; underlying ecosystem sensitivity to acid deposition, biogeochemical, atmospheric 

and deposition processes; and characterization of ecosystem services. While several processes 

are imbedded in the AAPI equation introduced in Chapter 5, the level of the AAPI, as in all 

NAAQS, is to include consideration of information on uncertainty and variability. 

Consequently, knowledge of the relative confidence and natural variability in the structural 

components of the AAPI are considered in staff conclusions on options for ranges of the level of 

the standard. This chapter is not intended to be a comprehensive treatment of all uncertainties 

that exist relative to the overall review of the standards, instead, it focused on those that are most 

relevant in evaluating choices regarding the AAPI form of the standard and options regarding the 

indicator, averaging time, and ranges of levels of the AAPI-based NOx and SOx standard. 

Uncertainty and sensitivity analyses are used to inform the relative confidence in the 

components and models that are used in defining the standard. Assessments of variability in the 

data used to determine parameters of the standard increases the level of understanding about the 

likelihood that alternative parameterizations of the standard will achieve targeted levels of 

protection when applied to sensitive ecosystems across the U.S. Assessments of the sensitivity 

of the overall AAPI to the components of the equation proposed to calculate the AAPI can help 

September, 2010 7-1 Draft – Do Not Quote or Cite 
demonstrate how important uncertainty and variability in those components are in assessing the 

protection of ecosystems provided by an AAPI standard. To evaluate the potential interactions 

between uncertain and/or variable AAPI components, a multifactor sensitivity analysis is also 

conducted. The ranges of component values evaluated in the multifactor sensitivity assessment 

are guided by individual variability and uncertainty analyses of specific components. In addition 

to informing considerations of the AAPI level, an additional objective of these “confidence” 

related analyses and discussions is to help guide research and data collection efforts intended to 

reduce uncertainty for future NAAQS reviews and implementation efforts. Spatial and 

temporal variability analyses of AAPI components are especially useful to inform monitoring 

network design, the spatial boundaries of acid sensitive regions, and averaging periods relevant 

to NAAQS implementation. 

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Use of confidence assessment results to inform setting of the AAPI level.

The analyses summarized in this chapter largely address parameters related to 

atmospheric/deposition and biogeochemical characterization, which are incorporated in the 

AAPI. General uncertainty related discussions are provided for elements in the standard setting 

process including dose-response relationships, effects, economic valuation metrics, as well as for 

several specific processes (e.g., organic nitrogen) which are directly represented in the AAPI. 

Confidence related information is not used to make decisions on inclusion or exclusion of a 

parameter in the AAPI model. Instead, we develop the overall approach using the best available 

information and most practical approaches, with an understanding of the level of confidence 

associated with both individual components and the overall resulting AAPI. The overarching 

principle in setting the level of the AAPI is consideration of the impact that associated NOx and 

SOx levels will have on the level of protection from adverse effects on public welfare. 

Information on uncertainty in different elements of the form of the standard and the scientific 

evidence informing those elements is used to evaluate our confidence that the standard has a high 

likelihood of protecting public welfare. One way of considering confidence levels is to 

identify components of the standard for which we have low confidence and adjust the level of the 

AAPI to ensure that, given the uncertainty in those components, we have a high likelihood of 

achieving a target level of protection. This approach to considering confidence levels places 

greater weight on ensuring a high likelihood of achieving protection, and places less weight on 

September, 2010 7-2 Draft – Do Not Quote or Cite 
concerns about providing more that requisite protection. Conversely, low confidence in 

components of the standard could also be considered to increase the likelihood that more than 

requisite protection could result, and consequently, the AAPI could be adjusted downward to 

account for lower confidence that the level of protection is requisite. To the extent that the 

available information suggests that a particular process or parameter creates a positive or 

negative bias (as opposed to broader uncertainty with no clear direction) would lead to 

recommending correspondingly higher or lower AAPI level. The Clean Air Act (CAA) 

language requires the secondary standards to be set to provide a requisite level of protection 

against known or anticipated effects, which can be interpreted to include effects which have 

large uncertainty but for which the expectation of adverse effects on public welfare exists with 

reasonable confidence. 

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Significant emphasis is placed on evaluations of CMAQ due to the unique role that 

atmospheric models hold in the linked AAPI system. The AAPI as currently formulated relies 

on CMAQ for both the initial characterization of reduced nitrogen deposition, and the deposition 

transformation ratios (TNOx and TSOx) which characterize the relationships between atmospheric 

concentrations of NOx and SOx and deposition of N and S. Included are interpretations of 

model evaluation results from the REA (EPA, 2009) as well as more recent results related to wet 

deposition and the treatment of ammonia deposition. Comparison of model results to 

observations provides a general sense of the confidence we have that the models capture the 

spatial, temporal and compositional texture of the relevant atmospheric and deposition species 

that drive the linked atmospheric-ecosystem processes. Both model evaluation results and 

assessments of spatial and temporal variability guide implementation strategies for monitoring 

network design and emission inventory improvement. Sensitivity of CMAQ derived deposition 

transformation ratios to changes in emissions, and treatment of chemistry [not yet completed] 

and variability over time provide insight into the stability of these parameters that are used in a 

relatively static manner in the AAPI, and into how well proposed averaging times capture the 

overall spatial and temporal trends in the parameters. 

We evaluate the sensitivity of critical load modeling components by comparing dynamic 

(MAGIC) and hybrid steady state model results, looking at terminal results of MAGIC. This 

approach was viewed as a test of the more reduced form approximations used in steady state 

September, 2010 7-3 Draft – Do Not Quote or Cite 
modeling relative to more sophisticated treatment in MAGIC. The MAGIC critical load 

simulations also provide information on the temporal trajectory of ANC, including the expected 

time necessary to reach a desired ANC, which can help inform the level of the AAPI, 

recognizing that there may be additional consideration given to reaching a target ANC within a 

specific timeframe, e.g. by 2030 or 2040. [[not yet completed] 

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For the purposes of this discussion, we characterize uncertainty regarding models and 

their outputs as referring to the lack of knowledge regarding both the actual values of model 

input variables (parameter uncertainty) and the model characterization of physical systems or 

relationships (model uncertainty). In any application, uncertainty is, ideally, reduced to the 

maximum extent possible, but significant uncertainty often remains. It can be reduced by 

improved measurement and improved model formulation. Model evaluation results provide 

some insight into the relative uncertainty associated with the ability of models to capture key 

environmental state characteristics. Confidence regarding the fundamental science supporting 

causal determinations about the effects of acid deposition, and the translation of those effects 

into ecosystem services and values is less amenable to quantification. As a result, these 

uncertainties are more difficult to explicity account for in development of the standards. In the 

case of the equation describing the AAPI, while the degree of uncertainty in some elements can 

be characterized, sometimes quantitatively, a formal uncertainty analysis using statistical 

sampling techniques (e.g., Monte Carlo simulation) to identify the relative and combined 

influences of parameter uncertainty was not performed. However, we did evaluate the sensitivity 

of the AAPI to its components using two related assessments: analysis of variance (ANOVA) 

and elasticity calculations. The results of these assessments are addressed in section 7.5. 

Sensitivity refers to the influence on modeled results due to perturbations in input 

variables or change of process formulations. Sensitivity analysis can provide a sense of how 

important different parameters and inputs might be to the outcomes of interest, e.g. the AAPI 

level, but cannot by themselves indicate how important specific parameters actually are, because 

they do not incorporate information on the range of parameter values or the likelihood associated 

with any specific parameter value. Sensitivity results in this PAD are intended to provide 

insight into the relative stability of the AAPI and associated NOx and SOx tradeoff curves and 

confidence in modeled parameterizations. Sensitivity analyses are especially useful in the 

September, 2010 7-4 Draft – Do Not Quote or Cite 
absence of observed data to challenge models. For example, the NOy and SOx transference 

ratios are a model construct that is difficult, if not impossible, to compare to observations. The 

sensitivity of these ratios to changing meteorology, emissions and chemical mechanism 

treatments is evaluated in reference to the stability of these ratios under changing conditions. 

Low sensitivity here implies that the choice to use long-term averages of modeled ratios is 

justified. Sensitivity analyses also are used to discern the relative influence (on AAPI results) of 

AAPI parameters. Toward that end, ANOVA and elasticity analyses were applied to determine 

the relative sensitivity of AAPI results associated with individual and combined AAPI 

parameters. 

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Variability refers to the heterogeneity in a population or variable of interest that is 

inherent and cannot be reduced through further data collection and research. In the context of 

the AAPI and trade-off curves, variability is considered in guiding the design of monitoring and 

modeling analyses supporting implementation activities. 

7.2 Uncertainty associated with ecosystem effects and dose – response relationships. 

Chapter 2 provides a brief summary of uncertainties based on the REA and is reproduced 

here to centralize all uncertainty discussions. 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 ANC and atmospheric deposition. 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 (REA 7.0): 

 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. 

 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. 

September, 2010 7-5 Draft – Do Not Quote or Cite 
1  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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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 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 was derived from multiple site calibrations. 

The 95% confidence interval for pre-acidification of lakes was an average of 15 μeq/L difference 

in ANC concentrations or 10% and 8 μeq/L or 5% for streams (REA 7.1.2). The use of the 

critical load model used 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). To propagate 

uncertainty in the model parameters, Monte Carlo methods were employed to develop an inverse 

function of exceedences. There is high confidence associated with the ecological effect function 

developed for aquatic acidification. In calculating the ANC function, the depositional load for N 

or S is fixed by the deposition of the other, so deposition for either will never be zero (Figure 

7.1-6 REA). 

Chapter 2 also reviews the basic evidence underlying effects on fish mortality, aquatic species 

diversity and more extended food web disruptions leading to adverse impacts on birds associated 

with aquatic acidification. There is high confidence associated with causality between 

acidification and these ecological effects. Also, there is extremely high confidence in the 

relationship between the ecological indicator, ANC, and the more direct chemical properties 

(lower pH and increased Al) associated with acidification. 

September, 2010 7-6 Draft – Do Not Quote or Cite 
1 7.3 Uncertainty in benefits estimates 

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Descriptions of the current provision of ecosystem services presented for each of the 

effect areas analyzed for this review followed by estimations of the damages incurred to selected 

services due to nitrogen and sulfur deposition. The current services are presented to give the 

reader a sense of the magnitude of the benefit the public receives from these ecosystems under 

current conditions. The data used in these descriptive passages is generally derived from

government (either federal or state) sources we are reasonably certain to be of the highest 

quality. Where monetary values are placed on the these services we have generally used widely 

cited studies, particularly meta analyses that provide an average value that smoothes the variation 

in WTP estimates. These estimates underestimate the total value of these services as they use 

benefit estimates for a marginal increase in these services. It is likely that the total benefits of 

these services are greater because their marginal value likely is lower than the average value. 

While reductions in sulfur and nitrogen emissions would increase the size of the benefits from

these services, for many of them it is unknown how significant the increase will be. 

The analyses of damages incurred are more uncertain and are limited to those areas where 

data and tools were available. Only some services were analyzed which in some cases meant 

that the results were limited to one or two services and in the case of terrestrial nutrient 

enrichment no services had sufficient data available to attempt an estimate of damage. This 

means that the estimates presented are a very small part of the total damage incurred due to 

deposition. 

Aquatic Acidification 

Recreational Fishing Model. 

The analysis of recreational fishing damages presented in Chapter 3 is subject to the 

assumptions necessary to perform the analysis. The original analysis performed for the REA 

was based on projecting future benefits of increased recreational fishing based on a complete 

cessation of all nitrogen and sulfur emissions. These decisions under or over estimate the current 

damages to public welfare incurred from nitrogen and sulfur deposition. The magnitude of the 

bias in results is unknown in either direction however the majority of the assumptions influence 

September, 2010 7-7 Draft – Do Not Quote or Cite 
the estimates downward. These include the use of emissions estimates that include projected 

decreases due to implementation of Title IV regulations in 2020. These emissions estimates are 

lower than current emissions and therefore lead to underestimation of damages. Because the 

models only value this improvement for New York residents (without accounting for out-of-state 

visitors) the damages are underestimates of the benefits of these improvements in the 

Adirondacks region. 

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The use of projected population in the REA analyses contributes to an overestimate of 

current damages since current population is smaller than future population. Further, these 

estimates are extrapolated from a 44 lake subset and applied to all Adirondack lakes. The 

representativeness of this sample is unknown. This analysis also does not account for any 

change in fishing demand (possible overestimate) and income (possible underestimate). 

Banzhaf, et al Benefits Transfer. 

The approach using the WTP estimates from the Banzhaf study is subject to the same 

uncertainties described above and some additional considerations. Specifically there is some 

uncertainty regarding which types of ecosystem services are reflected in the study’s estimates of 

the improvements in ecosystem services of reducing acidification, particularly provisioning and 

regulating services. The values likely include recreational fishing services, which means they 

cannot be added to the recreational fishing model results, and other cultural services including 

other recreation and nonuse services. The inclusion in the survey of other ecosystem changes 

(birds, trees, etc.) leads to an overestimation of WTP for remediation of lake acidification alone. 

Finally, assumptions were required to align the Banzhaf survey scenarios to the likely results of 

complete removal of all nitrogen and sulfur emissions. These are reasonably close but not exact 

and may not be applicable to another baseline. 

Conclusion. 

While these estimates are subject to uncertainty we are reasonably confident that they 

represent a good first-order approximation of the damages to recreational fishing due to nitrogen 

and sulfur deposition. Additionally it should be noted that the Banzhaf survey results represent a 

broader picture (though by no means complete) of the damages to ecosystem services in the 

Adirondacks. Finally, we would again like to emphasize that these estimates represent only a 

September, 2010 7-8 Draft – Do Not Quote or Cite 
small sample of the damages incurred to a broad range of ecosystem services affected and the 

areas of the nation where acidic deposition is an ongoing issue. 

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7.4 CMAQ Application and Evaluation 

7.4.1 Overview of CMAQ model application 

The CMAQ model is a comprehensive, peer-reviewed (Aiyyer et al., 2007), threedimensional grid-based Eulerian air quality model designed to simulate the formation and fate of 

gaseous and particle (i.e., particulate matter or PM) species, including ozone, oxidant precursors, 

and primary and secondary PM concentrations and deposition over urban, regional, and larger 

spatial scales (Dennis et al., 1996; U.S. EPA, 1999; Bryun and Schere, 2006). CMAQ is run for 

user-defined input sets of meteorological conditions and emissions. For this analysis, we are 

using predictions from several existing CMAQ runs. These runs include annual simulations for 

2002 using CMAQv4.6 and annual simulations for each of the years 2002 through 2005 using 

CMAQv4.7 (Foley et al., 2010). CMAQv4.6 was released by the U.S. Environmental Protection 

Agency’s (EPA’s) Office of Research and Development (ORD) in October 2007. CMAQv4.7 

along with an updated version of CMAQ’s meteorological preprocessor (MCIPv3.4, Otte and 

Pleim, 2010)1 were released in October 20082. The 2002 simulation with CMAQv4.6 was 

performed for both the Eastern and Western domains. The horizontal spatial resolution of the 

CMAQ grid cells in these domains is 12 x 12 km. The 2002 through 2005 simulations with 

CMAQv4.7 were performed for the eastern 12-km domain and for the continental United States 

domain, which has a grid resolution of 36 x 36 km. The CMAQv4.6 and v4.7 annual simulations 

feature year-specific meteorology, as well as year-specific emissions inventories for key source 

sectors, such as utilities, on-road vehicles, nonroad vehicles, wild fires, and natural biogenic 

sources. Emissions for other sectors of the inventory for each of the years modeled rely on 

inventories for 2002. Details on the development of emissions, meteorology, and other inputs to 

the 2002 CMAQv4.6 runs can be found in a separate report (U.S. EPA, 2008). Inputs for the 

CMAQv4.7 runs for 2002 through 2005 were derived using procedures similar to those for the 

CMAQv4.6 2002 runs. 

1 The scientific updates in CMAQ v4.7 and MCIP v3.4 can be found at the following web links: 

http://www.cmascenter.org/help/model_docs/cmaq/4.7/RELEASE_NOTES.txt 

http://www.cmascenter.org/help/model_docs/mcip/3.4/ReleaseNotes 

2 The differences in nitrogen and sulfur deposition in the case study areas between CMAQ v4.6 and v4.7 for 2002 

are small, as described in Chapter 3. 

September, 2010 7-9 Draft – Do Not Quote or Cite 
Additional details of the modeling domain, emissions and meteorological inputs are 

provided in EPA (2009; REA Appendices). 

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4 7.4.2 CMAQ Evaluation, Sensitivity and Variability Analyses

Past results. A variety of comparisons of modeled estimates to observations were 

included in the REA (EPA, 2009), and some of the highlights are summarized here in addition to 

new work on ammonia characterization and wet deposition. Readers are encouraged to review 

the earlier report. Ambient air concentrations and wet deposition observations are paired against 

modeled estimates. In contrast, dry deposition is always a modeled value, either derived from 

ambient or modeled ambient concentrations. Given the interest in relevant nitrogen and sulfur 

species, CASTNET observations were used extensively. Comparisons of modeled annual 

average total nitrate (sum of nitric acid and particulate nitrate), ammonium, sulfate, and sulfur 

dioxide to observations for the 2002 base year are provided in Figures 7-1 through 7-4. 

Normalized mean bias statistics for 2002-2005 base years are provided in Table 7-1. 

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CMAQ overpredicts SO2 and underpredicts SO4. Although model performance is good 

for total SOx, the inclusion of co-located SO2 and sulfate measurements required for future 

secondary NOx/SOx NAAQS comparisons will help diagnose issues with the model’s ability to 

partition these two species. CMAQ generally overpredicts total nitrate and slightly 

underpredicts ammonium and the model captures the monthly temporal patterns of sulfate, total 

nitrate and ammonium when all sites are aggregated (Figures 7-5 to 7-7). There are some basic 

incommensurabilities between model estimates and observations that complicate interpretation 

of model to observation comparisons, most notably the representation of space as a model 

represents a volume average of roughly 144 km2

, which depends on the time varying vertical 

depth of the lowest modeled layer. Most surface based observations rely on point sampling and 

the extent to which a point is representative of broader volume space varies with meteorology, 

distribution of emissions and surface characteristics. 

September, 2010 7-10 Draft – Do Not Quote or Cite 
Table 7-1. Normalized Mean Bias Statistics for Predicted and Observed Pollutant 

Concentration

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Pollutant 

Concentrations 

2002 2003 2004 2005 

SO2 45% 39% 47% 41% 

SO4

2- -13% -9% -13% -17% 

TNO3 22% 26% 22% 24% 

NH4

+ 4% 11% 7% 2% 

3 Ammonia. 

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Characterizing ammonia deposition is of increasing importance for assessing ecosystem 

responses to nitrogen deposition. Nitrogen deposition includes both deposition of NOy and 

deposition of reduced nitrogen (ammonia + ammonium). Ecosystem effects are due to total 

nitrogenand, as such, the recommended AAPI form of the standard directly includes reduced 

nitrogen to account for its role in using up an ecosystems ability to absorb nitrogen. The 

proposed form of the standard uses CMAQ modeled deposition of reduced nitrogen. 

Because of a shortage of routinely available ammonia observations, model evaluation 

studies have relied on ammonium measurements from CASTNET as the only routine source of 

reduced nitrogen observations. Clearly, the lack of ammonia observations must be addressed in 

future implementation scenarios as there is relatively greater uncertainty in characterizing 

ammonia relative to SOx and NOy. Recently, Dennis et al. (2011) explored the sensitivity of 

CMAQ (addressed below) to varying treatments affecting ammonia deposition velocity. As part 

of that study, CMAQ ammonia estimates were compared with ammonia observations at two sites 

in North Carolina (Figure 7-8). 

18 Wet deposition. 

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Modeled wet deposition in CMAQ is a function of the volume of predicted precipitation 

within a grid cell and the pollutant concentrations scavenged from the atmosphere during 

precipitation events. As a result, errors in modeled precipitation and in emission inputs can lead 

to significant bias and error in the wet deposition predictions compared to observed values. EPA 

September, 2010 7-11 Draft – Do Not Quote or Cite 
(Dennis and Foley, 2010) has corrected CMAQ wet deposition predictions by scaling the model 

output based on observation-based gridded precipitation data generated by the Parameterelevation Regressions on Independent Slopes Model (PRISM, 2004). The precipitation adjusted 

deposition fields are more highly correlated with observed values for all wet deposited nitrogen 

and sulfur species compared to the base model output (Figures 7-9). In addition, the adjusted 

fields are better able to capture the spatial heterogeneity of accumulated wet deposition due to 

orographic effects on precipitation amounts. 

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Adjusting the wet deposition values to account for over-predictions in the model 

precipitation inputs revealed compensating errors for nitrate and ammonium. The negative bias 

seen in these species after the precipitation adjustment is believed to be due to missing emissions 

sources. A second bias adjustment was performed for nitrate and ammonium based on observed 

levels at the NADP/NTN sites (Figure 7-10). The final adjusted spatial fields of annual total wet 

deposition values are more consistent with observed wet deposition values. Ongoing studies 

suggest that much of this bias can be reduced in the Eastern half of the US by including nitrogen 

oxide produced by lightning and accounting for the bi-directional flux of ammonia. Once these 

model improvements are incorporated in CMAQ a second bias adjustment may not be needed in 

the East. 

 7.4.3 Variability and sensitivity of CMAQ generated components. 

7.4.3.1 Ambient Concentration to Deposition Transformation ratios. 

The deposition transference ratios3 introduced in Chapter 5 are referenced as TSOx and 

TNOy, to distinguish these parameters from an exact linkage to deposition velocity, which is 

uniquely associated with individual atmospheric species. Deposition transference ratios are 

defined as the annual wet and dry deposition of all oxidized species (NOy for TNOy, SO2 plus SO4

for TSOx) divided by the average annual concentration of NOy, for TNOy, or SO4 plus SO2, for 

TSOx. The units for TNOy and TSOx are distance/time. Deposition transference ratios provide a 

mechanism to associate ambient concentrations to deposition loads and to determine if an area’s 

3 In the first draft of the Policy Assessment, the deposition transformation ratios were labeled VNOx and VSOx. For 

this draft, based on recommendations from CASAC, we have renamed these ratios TNOx and TSOx. 

September, 2010 7-12 Draft – Do Not Quote or Cite 
air concentrations of NOy and SOx meet a NAAQS level using the AAPI form. A deposition 

transformation ratio is an aggregate representation of the deposition process generated through 

modeling which does not lend itself to a traditional analysis relating observations and 

predictions. Furthermore, there is an implicit assumption that the response of deposition 

transformation ratios to changes in meteorology and emissions is relatively stiff, as these ratios 

are an attribute of the system that channels ambient air response associated with decreases in 

emissions of NOx and SOx to changes in deposition. The stiffness of the deposition transference 

ratios would suggest that the relationship between ambient concentrations and deposition is 

strictly a constant proportion, not impacted by the mixture and level of emissions or by changes 

in meteorology. To better understand the implications of this assumption, we investigated the 

relative variability of the modeled deposition transformation ratios across time and space, and the 

stability of the ratios relative to emissions and meteorological inputs was conducted to guide 

EPA in determining how uncertainties in this parameter may eventually impact AAPI related 

calculations. 

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15 7.4.3.2 Spatial and Interannual Variation of TS and TN. 

Generally small spatial and inter-annual variability exist in the deposition transformation 

ratios for the 2002 -2005 model years (Figure 7-11). The inter-annual variability, calculated at 

the grid cell level, as measured by the median coefficient of variation is around 10% and the 

absolute values of the ratios remain stable (Figure 7-12), suggesting that year to year changes in 

meteorology have minimal impact on the ratios. Spatial homogeneity of deposition 

transformation ratios within the two acid sensitive areas we evaluated in the REA (Adirondacks 

and Shenendoah) (Figure 7-11) is consistent with a relatively homogeneous ambient 

concentration environment overlaid upon a landscape of similar vegetation and surface 

conditions. Such spatial homogeneity within case study areas provides confidence that an area 

wide application AAPI will not be strongly dependent on the exact boundaries chosen to define 

an acid sensitive area. 

September, 2010 7-13 Draft – Do Not Quote or Cite 
1 7.4.3.3 TSOx and TNOy Sensitivity to emission changes. 

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The response of TSOx and TNOy to emission changes was explored by analyzing available 

base case 2005 and 2030 CMAQ simulations. The 2030 case reflected expected changes in 

emissions associated with simulated implementation of a variety of national rules and represents 

Eastern U.S. domain wide NOx and SOx emission reductions of 48% and 40%, respectively. 

Median changes in deposition transference ratios tended to be around zero (figure 7-12), with the 

Adirondack region exhibiting slightly higher response than the Shenandoah region and 

remainder of the Eastern U.S. domain. 

7.4.3.4 TSOx and TNOy sensitivity to different chemical mechanisms. 

10 [Intended to be added in the final PA] 

September, 2010 7-14 Draft – Do Not Quote or Cite 
1 7.4.3.5 Ammonia sensitivity. 

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The role of NHx deposition is incorporated in the AAPI expression as a parameter that 

influences the level of allowable concentrations of NOy and SOx, due to its role as part of the 

total reactive nitrogen budget which affects acidification.. Characterizing ammonia deposition 

is challenging due to the variety of surface and vegetation types that influence ammonia dry 

deposition velocities as well the potential for bi-directional flux of ammonia. In addition, 

ammonia emission estimates remain relatively more uncertain than emissions of NOx and SO2

given the complexity of meteorology and agricultural practices that influence the spatial and 

temporal patterns of ammonia releases. An exploration of the sensitivity of ammonia to three 

different treatments of deposition processes in CMAQ was performed by EPA (Dennis et al., 

2010) to test the inclusion of a bi-directional NH3 flux algorithm and elucidate the relative 

importance associated with advection, deposition and chemical transformation on ammonia 

patterns. These treatments included a (1) base case of current CMAQ treatment using existing 

ammonia deposition velocity schemes and uni-directional deposition, (2) modified the base case 

by replacing ammonia deposition velocity calculations with SO2 deposition velocities (SO2

interacts with surfaces and vegetation similarly to NH3, but with reduced velocity) as a lower 

bound and (3) introducing a bi-directional flux algorithm to the base case (retaining NH3 

deposition velocities). Based on modeled process analysis that delineates the effects of 

deposition, chemical transformation and advection (horizontal and vertical) on emitted ammonia, 

the results (Figure 7-13) suggest that ammonia patterns, especially when a bi-directional flux 

process is incorporated, are more indicative of a transported pollutant where emissions influence 

can span hundreds of kilometers, markedly different from some earlier perspectives where 

ammonia often was thought of as near source phenomenon due to high deposition velocities. 

The process analysis illustrates the importance of vertical advection which enables the movement 

of ammonia into traditional mesoscale flow patterns. The effect is enhanced by the 

reintroduction of deposited ammonia through bi-directional flux into the ambient environment. 

From a monitoring perspective, a design that addresses the regional characterization of 

NOy and SOx would be consistent with characterizing NHx. Not only would ammonia and 

ammonium measurements be useful for estimating dry deposition through deposition modeling 

approaches such as those used in CASTNET, but they would serve as important diagnostic data 

September, 2010 7-15 Draft – Do Not Quote or Cite 
to continually assess the effectiveness of NHx deposition processes in models like CMAQ. This 

is especially important as we recognize a large uncertainty in the bi-directional formulation 

associated with the estimation of Γ, the emissions potential due to the existence of compensation 

points. Nonetheless, we can learn much about the NH3 budget in spite of these uncertainties. 

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High priority research is ongoing to improve the bi-directional parameterization and the 

estimates of the leaf and soil gammas across different cropping regions and throughout the year. 

We are developing a software tool to estimate the soil Γ associated with fertilizer application. 

When we have a spatially and temporally varying Γg, we will investigate the emissions budgets 

for fertilized fields, as well as reexamine the animal operation emission budgets, as this will be 

of interest. Work to examine the seasonality of single cell budgets and their range of influence is 

continuing. Current and future CMAQ applications to ecosystem deposition will incorporate bidirectional flux treatment of ammonia. 

7.4.3.6 CMAQ uncertainties and the AAPI 

The AAPI relies on CMAQ for the sulfur and nitrogen transference ratios and NHx

deposition. The model evaluation results, including the ammonia and wet precipitation 

treatments, reflect a continual process of model improvement designed to ingest the latest 

science within a framework links a myriad of atmospheric and surface processes across multiple 

pollutant species. While this document focuses in on the more direct processes affecting N and 

S deposition, these modifications are incorporated with the philosophy that the best science is 

being adopted and they in turn support the overall improvement of the models’ treatment of all 

processes. The inclusion of better chemistry and physics of a particular process acts as an 

internal diagnostic tool for other processes that are linked throughout the model framework 

through basic conservation of mass principles. With respect to the AAPI, the CMAQ model 

must be relied on to provide the spatial flexibility attendant with a national standard. As the 

model continually adopts the best science, confidence in relevant CMAQ generated AAPI 

parameters is raised for both near term and future scenarios. 

September, 2010 7-16 Draft – Do Not Quote or Cite 
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Figure 7-1 2002 CMAQv4.6 annual average SO2 predicted concentrations versus 

observations at CASTNet sites in the eastern domain (note, units are in 

actual mass for SO2, including oxygen). 

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Figure 7-2 2002 CMAQv4.6 annual average SO2-

4 predicted concentrations versus 

observations at CASTNet sites in the eastern domain (note, units are in 

actual mass for SO4, including oxygen).

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Figure 7-3 2002 CMAQv4.6 annual average TNO3 predicted concentrations versus 

observations at CASTNet sites in the eastern domain (note, units are in 

actual mass for NO3, including oxygen). 

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Figure 7-4. 2002 CMAQv4.6 annual average NH4

+

 predicted concentrations versus 

observations at CASTNet sites in the eastern domain (note, units are in 

actual mass for NH4, including hydrogen). 

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Figure 7-5. 2002–2005 Domain-wide average SO4

2- predicted concentrations and 

observations by month at CASTNet Sites in the eastern domain(note, units 

are in actual mass for SO4, including oxygen) . 

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Figure 7-6. 2002–2005 Domain-wide average TNO3 predicted concentrations and 

observations by month at CASTNet sites in the eastern domain (note, units 

are in actual mass for NO3, including oxygen). 

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Figure 7-7 2002–2005 Domain-wide average NH4

+

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observations by month at CASTNet sites in the eastern domain (note, units 

are in actual mass for NH4, including hydrogen). 

September, 2010 7-23 Draft – Do Not Quote or Cite 
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Kenansville Ammonia July 2004

12-hour Averages: 6am-6pm

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184 186 188 190 192 194 196 198 200 202 204 206 208 210 212 214

Julian Day: TickMark at Midnight (July 2004)

NH3 (ug/m3)

Observations CMAQ

Millbrook (Raleigh) Ammonia July 2004

12-hour Averages: 6am-6pm

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183.5 185.5 187.5 189.5 191.5 193.5 195.5 197.5 199.5 201.5 203.5 205.5 207.5 209.5 211.5 213.5

Julian Day: TickMark at Midnight (July 2004)

NH3 (ug/m3)

Observations CMAQ 3

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Kenansville Ammonia August 2004

12 hour averages: 6am-6pm

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Julian Day: TickMark at Midnight (August 2004)

NH3 (ug/m3)

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Millbrook (Raleigh) Ammonia August 2004

12 hour averages: 6am-6pm

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Figure 7-8. Comparison of CMAQ predictions and measurements for 12-hour (6am6pm) average NH3 concentrations, with a monitoring cycle of 4 days on and 

4days off, at a high emission site (Kenansville) and a low emission urban site 

(Raleigh) in North Carolina compared to CMAQ for July 2004 (top) and 

August 2004 (bottom), from Dennis et al., 2010 (note, units are in actual mass 

for NH3, including hydrogen). 

September, 2010 7-24 Draft – Do Not Quote or Cite 
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Figure 7-9 Unadjusted (left) and PRISM (right) adjusted CMAQ annual wet deposited 

sulfate for 2002 (note, units are in actual mass for SO4, including oxygen). 

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Figure 7-10 Unadjusted (left) and PRISM and bias (right) adjusted CMAQ annual wet 

deposition of nitrate (top) and ammonium (bottom) (note, units are in actual 

mass for NH4 and NH3, including hydrogen). 

September, 2010 7-26 Draft – Do Not Quote or Cite 
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Figure 7-11 Spatial and interannual variability of inverse deposition transference ratios, 

1/TSOx and 1/TNOy, for Adirondack (top) and Shenandoah case study areas. 

September, 2010 7-27 Draft – Do Not Quote or Cite 
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Figure 7-12. Summary of inter-annual and emissions sensitivity variability of sulfur and 

nitrogen deposition transference ratios. 

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Figure 7-13. Cumulative regional NH3 budget of advection, wet- and dry deposition, 

calculated for an expanding box starting at the high-emitting Sampson 

County NC cell (from Dennis et al, 2010) 

September, 2010 7-29 Draft – Do Not Quote or Cite 
1 7.5. Sensitivity of AAPI to component parameters 

2 An elasticity analysis was applied to investigate sensitivity of the AAPI to its components 

(Appendix A). The means, medians and quartiles of the AAPI component variables were based 

on the range variable values across ecoregions that overlapped with the CMAQ domains. 

Elasticities measure the percent change in the AAPI for a 1% change in the AAPI parameters: Q, 

Neco, NHx, , TNOy, TSOx, NOy, and (SO2 + SO4). 

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 ( ) 1 11

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0 SOSOTNOyT Q

NHx

Q

Neco  BC

Q

AAPI 7  NOy SOx 

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The elasticity results identified significance for all the AAPI parameters. Detailed 

results are provided in Appendix A. Base cation weathering, , and hydraulic flow rate, Q, 

exerted strong influence on AAPI, an expected result given the explicit dependency evident in 

the AAPI expression. The transference rations for NOy (TNOy) and SOx (TSOx) exhibited 

relatively less influence on AAPI calculations than all other parameters when evaluated at means 

of the variables. However, in some locations, when evaluated at other values of the variables, 

AAPI can be more sensitive to the deposition transformation ratios. 

* BC0

These results suggest focusing on the uncertainties in the non-atmospheric inputs, 

including base cation weathering and runoff rates, and the implications of those uncertainties in 

setting an AAPI that will have a high likelihood of providing the targeted level of protection. 

An analysis of variance (ANOVA) analysis of the AAPI parameters will be added to the final 

PA. 

7.6 Uncertainty in Critical Load and ANC modeling 

7.6.1 MAGIC modeling 

An extensive uncertainty analysis of the MAGIC model was conducted as part of the 

REA, and documented in Appendix 4 of the REA. This uncertainty analysis included 

comparison of MAGIC outputs with observed water chemistry and ANC values. The 

uncertainty analysis also included an approach for generating confidence intervals for predicted 

ANC, using ensembles of model results based on alternative model calibration methods. 

September, 2010 7-30 Draft – Do Not Quote or Cite 
The model performance comparisons documented in Appendix 4 of the REA show close 

correspondence between simulated and observed annual average surface water SO4, NO3, and 

ANC during the model calibration period for 44 lakes in the Adirondacks Case Study Area and 

60 streams in the Shenendoah Case Study Area. These comparisons are reproduced in Figures 7-

14 and 7-15. Comparisons in the ability of MAGIC to reproduce the temporal pattern of ANC 

for individual lakes was also assessed, and the model does reasonably well at matching the 

pattern of ANC, although the fit is not as good as during the model calibration period. 

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The estimated confidence bounds on predicted ANC suggest that the 95 percent upper 

confidence bound is on average 10 percent higher in lakes, and 5 percent higher in streams. This 

suggests relatively low uncertainty introduced by the MAGIC modeling assumptions. MAGIC 

modeling is used in developing the estimates of base cation weathering for comparison to the Ffactor approach described in Chapter 5. 

7.6.2 SSWC modeling 

As stated in Appendix 4 of the REA, uncertainties in some elements of the SSWC 

modeling are not well understood. The version of the SSWC model used here uses the F-factor 

approach to estimate the preindustrial base cation supply for a given catchment. While this 

approach has been widely applied in Canada and Europe, it has only been used in a few cases 

within the United States and its assumptions and parameters have not been fully evaluated for 

aquatic systems. The natural or preindustrial catchment supply of base cations (i.e. weathering 

rates) has the most influence on the critical load calculation and also has the largest uncertainty 

(Li and McNulty, 2007). The uncertainty and ability to accurately estimate this parameter has 

not fully been evaluated and its uncertainty is unknown. It is important to note that for the 

United States, there is only one study for surface waters critical loads that compared steady-state 

and dynamic models and different steady-state approaches (MAGIC and F-factor) (Holdren et al. 

1992) other than what is presented in Chapter 5. Holdren et al. 1992 compared critical loads 

calculated by the steady-state MAGIC and the SSWC F-factor model for lakes in the Northeast. 

In this study, steady-state MAGIC model yielded critical load values that show the same general 

trend and on average were 14 kg/(ha-yr) SO4 higher than those from the SSWC F-factor 

approach, which is consistent with results, presented in Chapter 5. The two models converge at 

September, 2010 7-31 Draft – Do Not Quote or Cite 
low critical, but diverge as the buffering potential for watersheds increase, as indicated by 

increasing critical loads. 

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The REA conducted an uncertainty assessment using Monte Carlo simulation methods to 

characterize the uncertainty in estimated critical loads using the SSWC, varying a number of 

important inputs including runoff rates, water chemistry variables, and acid deposition. The 

coefficients of variation (CV) for the estimated critical loads (standard deviation divided by the 

mean) were calculated for each lake in the study as a measure of relative uncertainty. 

The results of this uncertainty analysis show that the coefficients of variation are on 

average very low for target ANC values within the range we are recommending (20 to 50 μeq/L). 

The CVs for critical loads are only 5% and 9% for critical load limits of 20 and 50 μeq/L, 

respectively. Although the average CV is relatively small for the population of sites modeled, 

individual site CV can vary from 1% to 45%. This difference is due to the high degree of 

uncertainty in site specific parameters for particular sites. 

These analyses suggest that uncertainties introduced in the AAPI directly by the SSWC 

Factor model are likely to be moderate. Additional uncertainties are introduced by the 

generalization of the F-factor approach to estimate critical loads in locations where F-factors 

have not been developed. 

September, 2010 7-32 Draft – Do Not Quote or Cite 
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Figure 7-14 Simulated versus observed annual average surface water SO42-, NO3-, ANC, 

and pH during the model calibration period for each of the 44 lakes in the 

Adirondacks Case Study Area. The black line is the 1:1 line. (Source: 

reproduced from REA, Appendix 4, Figure 1.1-1) 

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Figure 7-15. Simulated versus observed annual average surface water SO42-, NO3-, ANC, 

and pH during the model calibration period for each of the 60 streams in the 

Shenandoah Case Study Area. The black line is the 1:1 line. (Source: 

reproduced from REA, Appendix 4, Figure 1.1-2) 

7.7. Modeling and Data Gaps (To be expanded in final PA)

Atmospheric and deposition processes. The previous section introduced two important 

enhancements regarding the treatment of wet precipitation and the bi-directional flux of 

ammonia. The interest in deposition of sulfur and nitrogen raises the potential importance of 

occult (cloud and fog related processes) deposition associated with mists and clouds, which may 

be particularly relevant for aquatic acidification of high elevation watersheds. Occult deposition 

currently is in the early stages of development within the CMAQ framework. 

September, 2010 7-34 Draft – Do Not Quote or Cite 
Lightning generated NOx emissions have been an active area of research over the last 

decade and approaches that incorporate lightning count data and estimated NOx generation based 

on satellite measurements and aircraft campaigns have been tested in modern air quality models, 

including CMAQ. Lightning NOx is hypothesized to increase upper tropospheric ozone levels 

and wet nitrogen precipitation, with relatively negligible impact on near surface ambient nitrogen 

patterns. It is anticipated that CMAQ will incorporate lightning NOx for EPA assessments in the 

2012 timeframe. 

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Interest in organic bound nitrogen has increased based on NADP measurements 

suggesting that organic nitrogen contributes as much as 30% of the total nitrogen in precipitation 

samples. Significant uncertainties regarding the origin and composition of organic nitrogen 

(Altieri et al., 2009) suggest a need for research to improve our understanding of organic 

nitrogen prior to developing parameterizations in air quality models. Questions regarding the 

the relative contribution of anthropogenic or natural sources as well as the effects of reentrainment from the surface require attention. 

Atmospheric Observations. Chapter 4 addresses the current state of atmospheric 

observations relative to the NOx/SOx secondary standard and Chapter 8 addresses preliminary 

recommendations for monitoring methods. This new standard poses measurement resource 

challenges as the current networks, with the exception of CASTNET and some National Park 

Service (NPS) efforts, are deficient in spatial coverage relevant to anticipated acid sensitive areas 

and the specific measurement needs related to NOy, speciated NOy and ammonia and 

ammonium. 

Source emissions. Anthropogenic emissions of nitrogen oxides (NO and NO2) and 

sulfur dioxide generally are believed to be well characterized as the major contributors of NOx 

and SO2 from energy generation and transportation sectors have a history of continuous 

improvements of emissions modeling as well as direct emission measurements for major power 

generating units. Greater uncertainty resides in natural emissions of NOx from lightning 

processes (discussed above) and soil and agricultural related phenomena. Both NOx and 

ammonia emissions are subject to re-emission after deposition as part of the complex cycling of 

nitrogen in soils and biota. Characterizing the variety of agricultural practices that impact both 

ammonia and NOx is complicated by the dispersed nature of agriculture processes as well as the 

September, 2010 7-35 Draft – Do Not Quote or Cite 
influence of various meteorological factors on relevant biogeochemical processes controlling 

transformation and removal of nitrogen species. 

1

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Ecosystem processes and surface water observations. [To be completed in final PA]

The critical load modeling approaches that produced the N/S deposition tradeoff curves require a 

variety of input data depending on the approach chosen. In general terms, the availability of 

watershed related deposition, soil and vegetation characteristics and surface water chemistry 

determine the approach taken. There is a relatively extensive source of data for critical load 

modeling in the Eastern U.S., as illustrated by the frequent reliance on the Adirondack and 

Shenandoah Case studies. For this PA, critical load estimates were developed for national level 

coverage, largely through SSWC modeling relying on surface water data. Several ecoregions 

included an extremely small sample size of critical load estimates that challenged the 

development of a national scale assessment of acid sensitive areas. A more thorough 

characterization of nitrogen retention, dissolved organic carbon, soil chemistry in all acid 

sensitive areas would lead to reduced uncertainties in applying the AAPI as well as future 

considerations for standards that incorporate terrestrial acidification and nutrient enrichment 

effects.

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7.8. Summary and Conclusions 

Uncertainty and natural variability exist in all of the components of the structure of the 

NOx and SOx standard introduced in this PA, and should be considered in establishing the level 

of the AAPI. A summary of the relative uncertainties of these components is provided in table 

7-1 (To be added). On balance, the confidence level in the information and processes associated 

with the linkages from ecological effects to atmospheric conditions through deposition and 

ecosystem modeling is very high. The considerable body of evidence is conclusive with regard 

to causality between aquatic acidification and biological and ecological effects. Confidence in 

the linkage associating aquatic acidification and ANC is extremely high, as the aquatic chemistry 

describing this relationship, while nonlinear, is relatively simple with regard to chemical species 

and reactions. The relationships between deposition and ANC, while complicated by a variety 

of biogeochemical and hydrological processes and data requirements within watersheds, are well 

established and the critical load models have been thoroughly vetted through the scientific 

community with a demonstrated level of successful evaluation. The linkages between ambient 

September, 2010 7-36 Draft – Do Not Quote or Cite 
September, 2010 7-37 Draft – Do Not Quote or Cite 

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concentrations of relevant species and deposition is best handled through air quality modeling 

systems like CMAQ. The relationship between concentrations and deposition loads is well 

characterized by these models, which are constrained by mass balance principles. While much 

of the physical and chemical processing that determine concentrations and consequent deposition 

is interwoven with numerous fundamental processes characterizing mass transport and 

atmospheric chemical oxidation, the science is relatively mature with years of applications and 

continued evolution of the models. The specific processes guiding nitrogen and sulfur 

chemistry and deposition are relatively simple. More challenging is the ability to parameterize 

processes at the air-surface interface which guide the estimation of deposition velocities and the 

re-emission of certain species, as well as many of the area wide natural processes and 

agricultural practices which influence emissions of oxidized and reduced forms of nitrogen. 

The variety of uncertainty, variability and sensitivity analyses included in this chapter 

have been conducted under the assumption that the basic model construct is solid, as discussed 

immediately above, and are used to inform conclusions regarding the level of the AAPI that 

incorporate consideration of uncertainty. These analyses are also useful in guiding 

implementation efforts related to future monitoring, emissions and model process improvements. 

 The influence of uncertainty on the level of the AAPI can be thought of as reducing or 

increasing relative stringency of the level to increase the likelihood that requisite protection of 

public welfare is provided. Throughout these discussions there is no apparent directional bias 

in the uncertainty regarding the biological, chemical and physical processes incorporated in the 

AAPI. From the perspective of valuation of ecosystem services, the estimates generally are 

believed to be biased low, meaning the values of reaching a target level of protection are 

underestimated. However, quantification of these values is perhaps the most uncertain of all 

aspects considered. Consequently, the level of the AAPI should be relatively high in a buffering 

context to account for the existence of uncertainties in several components. In addition to, but 

related to these uncertainties discussions, are considerations of time lag to reach a target level 

ANC due to ecosystem response dynamics, as well the uncertainties in the severity and 

prevalence of episodic events. Both of these considerations suggest support for an AAPI that is 

somewhat higher than the target ANC supported by the specific evidence and risk information. 
1 Table 7-2. Summary (Incomplete) of Qualitative Uncertainty Analysis of Key Modeling Elements in th

Potential influence of

uncertainty on risk

estimates

Source Description Direction Magnitude

Knowledge‐

Base

uncertainty*

Comm

(KB: knowledge base, INF: inf

estim

Major elements (and sub‐models) of the ecological effects to ambient concentration framework

Biological/ecosystemresponse to

acidification

Clear associations betweenaquatic acidification (pH,

elevated Al) and adverse

ecosystem effects (fishmortality, decreased species

diversity)

Both Low Low

Linkage between

direct acidification

species to ecological

indicator (ANC)

The relationships across ANC,

pH and dissolved Al are

controlled by well defined

aquatic equilibrium chemistry

Both Low Low

ANC is the preferred ecosystem 

relationship with pH and the d

the NOx/SOx standard.

Linkage between

ecological indictor

and adverse

ecological effects

Direct associations between

ANC and fish mortality and

species diversity

Both Low‐medium Low

Although the pH dependency o

directionally consistent. In ex

environments the relationship 

catchments are in relatively “le

natural conditions or extreme 

acid mine drainage). In sensit

relationship essentially is simila

direct acidification species and 

Deposition to ANClinkage throughCritical Loadmodeling

Both MAGIC and Steady State

critical load models are

applied to determine critical

load models. The Steady

State critical load model

formulation is used as the

foundation for deriving the

Both Low Low

The model formulation is well 

substantial amount of research 

the peer reviewed literature. 

associated with the availability 

model components.

September, 2010 7-38 Draft – 
September, 2010 7-39 Draft – 

Potential influence of

uncertainty on risk

estimates

Source Description Direction Magnitude

Knowledge‐

Base

uncertainty*

Comm

(KB: knowledge base, INF: inf

estim

AAPI equation.

Atmosphericconcentrations to

Deposition

Deposition is a direct function

of ambient concentration,

influenced by severalprocesses, and handled in this

PA through air quality

modeling.

Both Low Low

The model design is appropriat

temporal complexities that infl

well as the variety of atmosphe

not measured. Greater uncert

driving (e,g,, ammonia emissio

availability of observations to e

Sub‐components and data of individual models

DepositionTransference Ratios

CMAQ derived ratio of total

oxidized deposition toconcentration averaged over

one year

both low unknown

Transference ratios enable the 

deposition and the policy relev

and (SO2 + SO4). They are stric

cannot be evaluated in a tradit

context. The low sensitivity o

changes and inter annual mete

spatial variability indicate that 

stable.

Additional elements to be added (.BCo, NECO, Q, DOC, benefits, ambient obs, surface water obs, emissions, .....)
1 7.9 REFERENCES:

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Altieri, K.E., B.J. Turpin and S.B. Seitzinger, 2009, Composition of dissolved organic nitrogen 

in continental precipitation investigated by bu ultra-high resolution FT_ICR mass 

spectrometry, Environmental Science and Technology, 43, 18, 6950 – 6955. 

Foley, K.M., Roselle, S.J, Appel, K.W., Phave, P.V., Pleim, J.E., Otte, T.L., Mathur, R., Sarwar, 

G., Young, J.O., Gilliam, R.C., Nolte, C.G., Kelly, J.T., Gilliland, A.B., Bash, J.O. (2010) 

Incremental testing of the Community Multiscale Air Quality (CMAQ) modeling system 

version 4.7, Geosci. Model Dev., 3, 205-226. 

Otte, T.L. and Pleim, J.E. (2010) The Meteorology-Chemistry Interface Processor (MCIP) for 

the CMAQ modeling system, Geosci. Model Dev., 3, 243-256. 

PRISM Climate Group, Oregon State University, http://www.prismclimate.org, created 4 Feb 

2004 

Dennis, R. and K. Foley, 2009, Adapting CMAQ deposition fields for critical loads analyses, 

presented at 2009 NADP Confernce; manuscript in preparation 

Dennis, R., R. Mathur, J.E. Pleim and J.T. Walker, 2010, Fate of Ammonia Emissions at the 

Local to Regional Scale as Simulated by the Community Multiscale Air Quality Model, 

accepted by Atmospheric Pollution Research 

Aiyyer, A, Cohan, D., Russell, A., Stockwell, W., Tanrikulu, S., Vizuete, W., and Wilczak, J. 

2007. Final Report: Third Peer Review of the CMAQ Model. 

Byun, D.W., and Schere, K.L. 2006. Review of the Governing Equations, Computational 

Algorithms, and Other Components of the Models-3 Community Multiscale Air Quality 

(CMAQ) Modeling System. J. Applied Mechanics Reviews, 59 (2), 51–77. 

September, 2010 7-40 Draft – Do Not Quote or Cite 
1 8 AMBIENT AIR MONITORING 

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Ambient air measurements of nitrogen and sulfur species support implementation of this 

proposed NAAQS secondary standard and improve the information basis for subsequent reviews. 

These uses extend beyond the basic need to measure the proposed NAAQS indicators, NOY, SO2

and SO4, in approximate priority order: 

 Direct NAAQS comparisons. 

 Reduced nitrogen, NH3 and NH4, to evaluate CMAQ and other air quality models 

ability to characterize NHX deposition, a component of the AAPI expression. 

 Model and process improvement - In combination with NAAQS indicators and NHx, 

additional speciated NOY components including HNO3, PAN, NO2, and NO to 

continually assess air quality model behavior and associated deposition processes. 

 Subsequent NAAQS reviews of the secondary NOX/SOX standard as well as related 

primary and secondary standards reviews (ozone, NO2, SO2 and PM). 

 Accountability – Assessing the effectiveness of implemented programs addressing 

emission strategies to meet attainment of the proposed NAAQS using all noted 

measurements. 

20 8.1 What Are The Appropriate Ambient Air Indicators To Consider In Developing The 

Standards? 

The recommendation of NOY, SO2 and SO4 as the ambient air indicators for the proposed 

NOX/SOX standard was introduced in the first draft of the PAD and endorsed by CASAC 

(CASAC, 2010). Essentially, NOY, which is an aggregate of all reactive oxidized nitrogen 

compounds and the two sulfur species represent the oxidized ambient air species of relevance to 

the criteria pollutants NOX and SOX with potential to adversely affect acid-base balance in 

aquatic systems. Contributions of reduced nitrogen, which would not be part of the indicator for 

the standard under the approach suggested in this PAD, are provided by CMAQ. 

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One could consider using all NOY species as NAAQS indicators, requiring, for example, 

measurements of the dominant species: HNO3, particulate nitrate, true NO2, NO, and PAN. 

Conceptually, each species would be paired with a species - specific deposition velocity in the 

AAPI expression to transfer individual deposition values to ambient values. The attraction of 

using individual species would be the reliance on actual deposition velocities that have more 

physical meaning in comparison to model constructed transference ratios which aggregate dry 

and wet deposition and all nitrogen species. The transference ratio approach does retain the 

necessary conservation of mass which underlies virtually all parameterization schemes, but loses 

some degree of physical relevance due to use of modeled outputs – an admittedly unique 

construct. The major drawback of using individual species as NAAQS indicators is the lack of 

routinely available measurement techniques and an associated resource burden even if adequate 

techniques were available. Currently, technology for measuring true NO2, HNO3, and PAN 

generally is not available for routine network applications. In addition to this practical 

consideration, there is another important reason for using the aggregated transference ratios. 

Because the standard, and the explicit Clean Air Act authority, is based on ambient air there must 

be an effective link connecting deposition and concentrations of the criteria pollutants in the 

ambient air. The use of individual species conceptually allows for a more physically meaningful 

approach to characterize and calculate deposition. However, the transference ratios also enable 

incorporation of the contributions of wet precipitation in the ambient air indicators. There is no 

practical alternative that allows for the disentangling of wet deposition as a function of ambient 

air concentrations as that relationship is best addressed through the coupling of numerous 

meteorological and chemical processes imbedded in the air quality modeling platform. One 

could consider wet precipitation as a separate parameter and isolate dry precipitation in the AAPI 

equation. But doing so would lose the important connection between wet precipitation of 

nitrogen and the same emission sources responsible for dry deposition. 

Finally, one might advocate for direct measurements of dry deposition of individual NOY

species. Again, technologies simply are not ready for consideration in routine network 

applications. And, it is arguably practical to model dry deposition even if direct dry deposition 

measurement technologies were available. One reason for this is that there is significant spatial 

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heterogeneity in the factors (vegetation and surface type, micrometeorology) that define 

deposition velocity. Consequently, direct dry deposition measurements would have limited 

spatial representativeness in comparison to ambient air observations. The model conceptually 

accounts for the spatial variance, at the level of horizontal grid cell resolution, of the factors 

defining deposition velocity on a species by species basis. However, one also could reason that a 

well placed direct measurement of dry deposition is more realistic than a modeled result that 

relies on numerous assumptions. The development of technologies to measure direct dry 

deposition will benefit the diagnosis and improvement of process formulations in models. 

Consideration has been given to change the atmospheric indicator from NOY to total 

nitrate (the sum of nitric acid and particulate nitrate). The rationale for that approach is that a 

larger fraction of the deposited NOY is accounted for by total nitrate, which currently is 

measured in CASTNET with high confidence. One can reason adequately that nitrate may 

correlate well with total oxidized nitrogen deposition relative to NOY (as discussed in Chapter 4) , 

given the inherent noise associated with variable contributions of low deposition velocity species 

of ambient level significance (e.g., NO2). The disadvantages of using total nitrate as an 

indicator are that significant ambient mass with the potential for deposition is not captured, and 

NOY is a preferred measurement for model evaluation and accountability purposes. 

Accountability refers to assessing if emissions reductions have the intended consequences on 

ambient air and deposition levels in the context of, “Aare our emissions reductions strategies 

working as planned?” All three of these concerns relate to the benefit of closing mass balances in 

whatever environmental medium is being characterized. In addition, the use of nitrate alone 

would create an increased distancing from the listed criteria pollutant oxides of nitrogen as NOY

does include NO and NO2. 

26 8.2 Reactive Oxidized Nitrogen and Sulfur Species. 

NOY species. Air quality models and deposition models that use direct observations calculate 

‘deposition on a species by species basis to account for differences in deposition velocities. 

Consequently, the relative fractional contributions of individual NOY or SOX species to 

deposition or concentration is influenced by the differences in species deposition velocities. For 

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example, nitric acid with a high deposition velocity would exhibit a larger relative contribution 

to overall deposition compared to ambient concentrations in a particular area (Figure 8-1). The 

dominant ambient air NOY species are NO, NO2, HNO3, P-NO3 and PAN. Near source urban 

environments typically have a relatively higher fraction of NOx compared to the products of 

NOX reactions, nitrates and PAN, which are relatively more dominant in rural locations (Figures 

8-2 – 8-5). 

The differences in the relative patterns between ambient air and deposition on a species 

by species basis illustrate a number of challenges and considerations in developing a monitoring 

strategy. It is clear in the Adirondacks and Shenandoah areas that nitric acid is the most 

dominant contributing species from a deposition perspective (Figure 8-1), with significant 

contributions from particulate nitrate, PAN and NO2. The original source of emissions (NO 

accounts for 90-95% of all emitted NOX) provides very small (< 5%) contributions to ambient air 

and virtually nothing to deposition. The combination of nitric acid and particulate nitrate 

consistently contribute greater than 50% of the oxidized nitrogen dry deposition load, whereas 

PAN and NO2 contribute roughly 15-25% of the deposition load. These broad summary 

statements speak to some of the monitoring considerations addressed earlier, particularly the case 

for monitoring for total nitrate. However, caution should be exercised when considering not 

measuring a considerable fraction of the ambient NOY burden reflected in NO2 and PAN. 

Characterization of NO2 deposition is an area requiring further refinement especially considering 

that NO2 is a significant component of total oxidized nitrogen. Zhang et al. (2005) suggest that 

NO2 contributes up to 36% of dry NOY deposition in rural Eastern Canadian locations, and 

suggest, based on observational evidence (Figure 8-2), that in some locations NO2 deposition 

may be similar to nitric acid contributions. 

A sampling of co-located NOY species observations in rural Eastern Canada (Figure 8-2), 

particularly in Egbert, Ontario, illustrates the concern of the general assumption that NO2 may 

not contribute significantly to NOY deposition in rural locations. While it may be true that in 

general NO2 is of less of concern in rural areas relative to urban areas, that does not dismiss the 

potential for significant misrepresentation of total nitrogen budget in certain rural locations. 

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The results also raise the question of potential efficiencies gained from not cycling 

between NO and NOY analyses, which is the standard configuration in commercial NOy 

instruments, acknowledging the limited use of NO data in rural, acid sensitive environments 

(note that NO present in the air would still be captured in the NOY measurement). 

These examples are used to support the rationalization of using NOY as an appropriate 

atmospheric indicator in applying the AAPI. However, while it may be required to measure 

NOY for explicit AAPI calculations to determine compliance with a NOx/SOx standard, there 

should be additional measurements of true NO2, HNO3, p-NO3 and PAN to allow for diagnostic 

evaluations of both air quality models and the NOY measurement itself. This recommendation 

would leverage existing CASTNET filter pack (FP) observations necessary to capture particulate 

sulfate (discussed below) and therefore require the addition of true NO2 measurements and 

periodic sampling for PAN. 

Sulfur Species. Although sulfur dioxide and particulate sulfate contribute approximately 60 and 

40 %, respectively, to ambient SOx concentrations, sulfur dioxide is the dominant contributor to 

SOX deposition (Figure 8-6), which is consistent with CASTNET observational studies (Sickles 

and Shadwick, 2007). Measurement technology issues are not as complex for SOx as they are 

for NOY and individual NOY species. Issues related to particle size fraction and averaging 

period for SOx are discussed below. 

Reduced Nitrogen. 

 The AAPI does not include reduced nitrogen (ammonia gas and ammonium ion) as an 

ambient air indicator. However, reduced nitrogen deposition is an explicit AAPI component 

which is estimated through air quality modeling. As discussed in the Chapter 7, characterization 

of reduced nitrogen deposition processes is an active developmental area which would benefit 

markedly from NHX measurements in order to assess modeled predictions of ambient patterns of 

ammonia and ammonium. This need for monitoring ammonia in rural environments is further 

supported by emerging evidence that ammonia acts as a regionally dispersed species based on 

the inclusion of ammonia bi-directional flux in CMAQ simulations. (Dennis et al., 2010). 

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Monitoring method approaches under consideration for routine application typically are limited 

to time averaged filter and denuder technologies, including passive sampling approaches. 

8.3 What measurements would be used to characterize NOy and SOx ambient air 

concentrations for the purposes of the AAPI based standard? 

Ambient NOy, SO2 and particulate sulfate (SO4) concentrations would be used as the 

indicators in determining compliance with the standard. All of these indicators are measured in 

different places within the current routine monitoring networks (section 3.2). However, there are 

issues requiring resolution associated with Federal Reference or Equivalency Measurement 

(FRM/FEM) status of measurement techniques that to date have served as supplemental 

information, which will require resolution. A FRM for SO2 exists, but not for NOYor SO4. Only 

recently have NOY measurements, which historically were viewed as research venue 

measurements, been incorporated as “routine” observations, partly as a result of the NCore 

program. Acquiring FRM status may require better characterization of the conversion 

efficiencies, mass loss and clear guidance on operating and siting procedures. Particulate sulfate 

has been measured for several years in the IMPROVE, CASTNET and EPA CSN networks. The 

nation has over 500 24-hour average, every third day sulfate measurements produced by the 

PM2.5 speciation networks (IMPROVE and EPA CSN) and nearly 80 CASTNET sites that 

provide continuous weekly average samples of sulfate with an open inlet accommodating all 

particle sizes. With minor exceptions, the PM2.5 fraction generally accounts for over 80% of the 

ambient sulfate mass. Unfortunately, as particle size diameters increase beyond 2.5 μ, 

gravitational settling imparts greater influence resulting in substantially enhanced deposition 

velocities. Consequently, the sulfate mass in size fractions greater than 2.5 μ potentially 

provides correspondingly greater contribution to as much as 50% of dry sulfate deposition in 

certain locations (Grantz et al., 2003). 

 Sample collection period is not an issue for gaseous measurements of NOY and SO2 that 

operate continuously. However, consideration should be given to using the CASTNET FP for 

SO2 measurements as a resource saving option, assuming the FPs will be used for particulate 

sulfate. However, the availability of highly time resolved data will support the continual 

September, 2010 8-6 Draft – Do Not Quote or Cite 
evaluation of SO2 and sulfate balance in air quality modeling systems which is a critical 

underpinning for both human and ecosystem health assessments. Some concerns have been 

raised about the possibility of exclusion of coarse particles from NOY samplers operating at low 

flow conditions as well as potential difficulties of reducing organically bound and mineralized 

nitrate. Insight into conversion and capture efficiency characteristics will be advanced both by 

research catalyzed by the need to support this standard and through ongoing and future network 

operations. 

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8.4 What additional complementary measurements are recommended? 

 We recommend that there be 3-4 locations nationally, in airsheds with different 

atmospheric chemistries, that sample not only for the NAAQS indicator NOY but for the suite of 

major NOY species as well; HNO3, p-NO3, PAN, true NO2, and NO as discussed earlier. Not 

only is this important from a modeling and process diagnosis perspective, but it is especially 

useful in the introduction of new measurements that have a limited track record to provide 

insight into instrument performance. In the case of NOY, it is even more relevant since there 

effectively are no standards that explicitly challenge instrument accuracy given the highly 

variable nature of NOY species distribution and the instability associated with mixing NOY gases. 

This quality assurance issue is analogous to PM2.5 where aerosol standards are not available and 

measurement accuracy is judged against periodic challenges relative to a “gold standard” 

instrument. Reduced nitrogen measurements of ammonia and ammonium ion are recommended 

at all locations with FRM/FEM instruments based on the need to support the AAPI as discussed 

above. 

8.5 What sampling frequency would be required? 

The averaging time for the standard is likely to be an annual average, perhaps based on 3-

5 years of data collection to minimize the influence of interannual varaibility in meteorology, 

especially precipitation.. Conceptually, extended sampling periods no longer than one year 

would be adequate for the specific purposes of comparison to a standard. However, there are 

significant peripheral benefits relevant to improving the scientific foundation for subsequent 

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reviews and a variety of related air quality and deposition assessments to be gleaned from more 

highly time resolved data. In particular, the critical role of air quality models in deposition 

assessments implies value to be derived from measurements that support model evaluation and 

improvement. Many of the monitoring approaches that are used throughout the nation sample 

(or at least report out) on daily (PM2.5 chemical speciation), weekly (CASTNET) and hourly (all 

inorganic gases) periods. There is a tradeoff to consider in sampling period design. For 

example, the weekly CASTNET collection scheme covers all time periods throughout a year, but 

only provides weekly resolution that misses key temporal and episodic features valuable for 

diagnosing model behavior. The every third day, 24-hour sampling scheme used in IMPROVE 

and EPA speciation monitoring does provide more information for a specific day of interest yet 

misses 2/3 of all sampling periods. The missing sampling period generally is not a concern when 

aggregating upward to a longer term average value as the sample number adequately represents 

an aggregated mean value. Additionally, there is a benefit to leveraging existing networks which 

should be considered in sampling frequency recommendations. A possible starting point would 

be to assume gaseous oxidized species, NOY and SO2, are run continually all year reporting 

values every hour, consistent with current routine network operations. Sulfate sampling periods 

should coincide with either the chemical speciation network schedules or with CASTNET. 

There are advantages to coordinating with either network. Ammonia gas and ammonium ion 

present challenges in that they are not routinely sampled and analyzed for, and the combined 

quantity, NHX is of interest. Because NHX is of interest, some of the problems of volatile 

ammonia loss from filters may be mitigated. However, for model diagnostic purposes, 

delineation of both species at the highest temporal resolution is preferred. 

8.6 What are the spatial scale issues associated with monitoring for compliance, and 

how should these be addressed? 

The current observation network for NOY, NHX and SOX is very modest and includes a 

monitoring network infrastructure that is largely population oriented with the exception of 

CASTNET and IMPROVE. While there is platform and access infrastructure support provided 

by CASTNET, NADP and IMPROVE, those locations by themselves are not likely to provide 

the needed spatial coverage to address acid sensitive watersheds across the United States. 

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Ambient monitoring at every watershed will not be required given the reality of resource 

constraints and the relative spatial homogeneity of air concentrations that are averaged over 

annual time periods and within ‘acid sensitive” areas. The spatial monitoring requirements will 

be associated with the determination of acid sensitive areas, which is discussed at length in 

Chapter 5. The number of sites per area will be addressed in rule development and general 

guidance based on an understanding of the spatial variability of NOY, NHX, sulfate and SO2 

combined with resource allocations will help inform those decisions. 

Critical load models applied for the purposes of this standard would be based on annual 

averages, which effectively serves to dampen much of the spatial variability. Furthermore, the 

development of an area-wide depositional load tradeoff curve implies focus on region wide 

characterization. Toward that end, CMAQ concentration fields will provide insight into the 

likely spatial representativeness of monitors leading to efficient application of monitoring 

resources. For example, the CMAQ based spatial coefficient of variation (standard 

deviation/mean) of oxidized nitrogen in the Adirondacks was 1.46%. Improved dry deposition 

estimates will result from enhancements of ambient monitoring addressing the N/S secondary 

standards as each additional location could serves a similar role that existing CASTNET sites 

provide in estimating dry deposition. 

8.7 What specific monitoring methods would be used? 

Federal reference and/or equivalent methods (FRM/FEM) are presently available only for 

SO2. Particulate SO4 is measured at over 500 sites nationally, and there is a general consensus 

that methods available are reliable and provide consistent data. NOY measurement is in a 

transition period from largely being viewed as a research level measurement to now being 

deployed as a routine measurement in EPA’s national 75 site NCORE network. The general 

consensus on NOY measurement is that the methodology is sound and applicable for 

routine/regulatory use, but there does not exist a well defined understanding of the quality of 

NOY data. Inorganic dry nitrate (nitric acid and particulate nitrate) is measured routinely in the 

CASTNET network with filter packs (FP). 

September, 2010 8-9 Draft – Do Not Quote or Cite 
One of the challenges associated with specifying performance attributes for p-SO4 and 

NOY is the lack of specific challenge standards. For example, instruments measuring discrete 

gases such as ozone or nitrogen oxide can be challenged by comparing an instrument’s reading 

when measuring known concentrations of gases which are readily provided for single gas 

concentrations. Particle standards are not available. NOY performance typically is challenged by 

known mixtures of NO2, and occasionally with N-propyl nitrate, which only addresses part of the 

spectrum of nitrogen species in an NOY mix. Consequently, instrument performance in EPA’s 

national networks for aerosol mass is quantified in terms of bias and precision relative to a colocated “performance evaluation” instrument. There is no comparable program in place for pSO4 or NOY. 

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p-SO4. The routinely operating methodology for particulate sulfate (p-SO4) is based on 

an integrated (i.e., time averaged over several hours or days) sample collection on a Teflon filter 

followed by ion chromatography (IC) detection in the laboratory. Two major variations of this 

approach are applied in the PM2.5 speciation (exclusion of particles larger than 2.5 μ and 24-hour 

collection typically every third day) and CASTNET (weekly average integrated sampling all year 

with an open inlet to include all size fractions). There are additional variations related to inlet 

design and flow characteristics of PM2.5 speciation samplers in which two designs are prevalent 

in the networks: (IMPROVE and EPA CSN SASS samplers). These variations are considered 

minor as sulfate species (dominated by ammonium sulfate) typically are not subject to major 

sampling artifacts associated with volatilization or condensation. The difference in inlets (open 

vs. 2.5 μ) is perceived by some as not an issue of concern as 80 - 90 % of the PM sulfate mass is 

distributed in size fractions less than 2.5 μ. However, the higher deposition velocities associated 

with larger diameter particles argue for including all size fractions as discussed above. 

Continuously operating in-situ sulfate instruments that allow for hourly, or less, data reporting 

are available. However, the limited deployment (less than 20 sites nationally) of these 

instruments combined with the 2.5 μ inlet cutoff configuration preclude consideration at this 

time. 

The CASTNET FP offers three important attributes: a history of high quality data, 

existing infrastructure and network to build on and an open inlet to capture the full range of 

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particle diameters. EPA intends to develop FRM status for this method. A significant 

additional advantages of using the FP method will be the availability of important co-measured 

species (e.g., SO2, total nitrate, ammonium). While EPA will expedite the certification process 

for the CASTNET FP, in the future consideration should be given to other available methods to 

more efficiently leverage network assets. For example, the SASS sampler potentially would 

accommodate ammonia gas and ammonium ion measurements, as well as other standard 

chemical speciation parameters depending on the configuration of this multi channel system. 

Continuous sulfate measurements would be extremely useful for model evaluation, especially 

considering the availability of continuous SO2 data that would be required as part of the NAAQS 

indicators. A performance based approach to meet equivalency requirements, given the variety 

of sulfate measurement approaches and well vetted and accurate analytical procedures. 

SO2. Sulfur dioxide is a NAAQS pollutant and a FRM is available. See 75 FR at 35554-

56 and 35593-95 (June 22, 2010) (adopting a second FRM for SO2). As part of the NCore 

network development effort, trace gas SO2 analyzers capable of sub ppb resolution became 

commercially available and are the preferred instruments for implementation in rural locations. 

As discussed above, the near continuous data output of gaseous analyzers is desired for 

peripheral support of model evaluation. Nevertheless, the convenience and resource savings 

associated with the CASTNET FP suggest that Federal Equivalency Method (FEM) status should 

be incorporated in concert with the sulfate certification process. 

NOY. In principle, measured NOY is based on catalytic conversion of all oxidized 

species to NO followed by chemiluminescence NO detection. While there are caveats associated 

with instrument conversion efficiency and possible inlet losses, the technique is considered 

adequate and routinely operational. Approximately 25 sites (out of a planned 75) in EPA’s 

NCORE network are operating NOY instruments, and an additional 5-1 10 sites are operated in 

SEARCH, CASTNET and other programs. NOY measurements are nearly continuous, reporting 

at hourly intervals providing far greater temporal information compared to filter or denuder 

based methods. FRM certificaton for NOY presents more considerable challenges given the 

limited history of routinely operating instruments. The process EPA is pursuing for certification 

status for NOY will be addressed in the final PA. 

September, 2010 8-11 Draft – Do Not Quote or Cite 
1 8.8 REFERENCES

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CASAC, 2010, Letter to Administrator Jackson on Review of the Policy Assessment of the 

Secondary National Ambient Air Quality Standard for NOx and SOx : First Draft., 

Grantz D.A., J.H.B. Gardner and D.W. Johnson, 2003, Ecological effects of particulate matter, 

Environment International, 29 (2003) 213 -239. 

NARSTO, 2010, Technical Challenges of Multipollutant Air Quality Management; Hidy, G.M., 

J.R. Brook, K.L. Demerjian, L.T. Molina, W.T. Pennell, and R.D. Scheffe (eds). 

Springer, Dordrecht, 2010. 

Sickles, J.E. and D.S. Shadwick, 2007, Changes in air quality and atmospheric deposition in the 

Eastern United States: 1990- 2004, JGR, 112, D17301 

Zhang, L., J.R. Broo, R. Vet, C. Mihele, M. Straw, J.M. O’Brien and S. Iqbal, 2005, Estimation 

of contributions of NO2 and PAN to total atmospheric deposition of oxidized nitrogen 

across Eastern Canada. Atmospheric Environment 39 (2005) 7030-7043. 

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Figure 8-1 Annual 2002 – 2004 CMAQ derived annual average fraction of ambient 

concentrations (above ) and deposition (below) of individual NOy species 

delineated by the Adirondack and Shenandoah case study areas and the 

remainder of the Eastern U.S. domain. 

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Figure 8-2 Examples of the Relative Abundance of Several NOy Species Measured at Two 

Rural Southeastern Canadian Sites as a Fraction of the Total Measured NOy

Concentration -- Kejimkujik, NS, (top) and Egbert, ON, (bottom) during 2003. 

Although both sites are in rural locations, the Kejimkujik, NS site represents more 

aged air masses as it lies considerably further downwind from major sources of 

NOx relative to the Egbert site. (Source: NARSTO, 2010) 

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Figure 8-3 Annual average fraction of NOy ambient air contributed by NO2 based on 

2005 CMAQ Eastern U.S. simulation at 12 km grid cell resolution. 

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Figure 8-4 Annual average fraction of NOy ambient air contributed by HNO3 based on 

2005 CMAQ Eastern U.S. simulation at 12 km grid cell resolution. 

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Figure 8-5 Annual average fraction of NOy ambient air contributed by PAN based on 

2005 CMAQ Eastern U.S. simulation at 12 km grid cell resolution. 

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Figure 8-6 Annual 2002 – 2004 CMAQ derived annual average fraction of ambient 

concentrations (above ) and deposition (below) of individual SOx species 

delineated by the Adirondack and Shenandoah case study areas and the 

remainder of the Eastern U.S. domain. 

September, 2010 8-18 Draft – Do Not Quote or Cite 
1 9 INITIAL CONCLUSIONS 

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Staff initial conclusions on the elements of the secondary NOX and SOX standards for the 

Administrator’s consideration in making decisions on the secondary NOX and SOX standards are 

summarized below, together with supporting conclusions from previous chapters. We recognize 

that selecting from among alternative policy options will necessarily reflect consideration of 

qualitative and quantitative uncertainties inherent in the relevant evidence and in the assumptions 

of the quantitative exposure and risk assessments. Any such standard should protect public 

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

pollutant(s) in the ambient air, “whether caused by transformation, conversion, or combination 

with other air pollutants.” CAA § 302(h). . In providing these options for consideration, we are 

mindful that the Act requires standards that, in the judgment of the Administrator, are requisite to 

protect public welfare. The standards are to be neither more nor less stringent than necessary. 

Our focus in this review on ecosystems that are both sensitive to acidification and responsive to 

atmospheric acid deposition is intended to ensure that the resulting standards are appropriately 

protective and not more protective than necessary in ecosystems that are not adversely affected 

by acid deposition. 

To evaluate whether the current secondary NAAQS is adequate or whether consideration 

of revisions is appropriate, the conclusions and options for the Administrator to consider in this 

review are based on effects-, exposure- and risk-based considerations. The exposure and risk 

assessments reflect the availability of new tools, assessment methods, and a larger and more 

diverse body of evidence than was available in the last reviews. We have taken a weight of 

evidence approach that evaluates information across the variety of research areas described in the 

ISA and in addition includes assessments of air quality, exposures, and qualitative and 

quantitative risks associated with alternative air quality scenarios. 

Staff notes that since the last review, additional policy-relevant developments have 

occurred that may also warrant consideration by the Administrator when making decisions about 

what is requisite to protect public welfare. The NRC report (described in Chapter 5) states: 

“Whatever the reason that led EPA to use identical primary and secondary NAAQS in the past, it 

is becoming increasingly evident that a new approach will be needed in the future. There is 

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growing evidence that the current forms of the NAAQS are not providing adequate protection to 

sensitive ecosystems and crops” (NRC, 2004). 

The last review raised the following key issues as a rationale for not setting a separate 

standard for NOX to protect against acidification and nutrient enrichment effects in sensitive 

ecosystems: 

1) Lack of enough consistent information to support a revision of the current secondary 

standard to protect these aquatic systems. 

2) Lack of adequate quantitative evidence on the relationship between deposition rates 

and environmental impacts 

3) Significant uncertainties with regard to the long-term role of nitrogen deposition in 

surface water acidity and with regard to the quantification of the magnitude and timing of the 

relationship between atmospheric deposition and the appearance of nitrogen in surface water. 

In this current review, staff concludes that important new information has become 

available since the last review that supports revising the current NOX and SOX standards. 

Specifically, the ISA has concluded that there are causal relationships between NOX and SOX

acidifying deposition and effects on aquatic and terrestrial ecosystems, and the ISA and REA 

provide substantial quantitative evidence of effects occurring in locations that meet the current 

NO2 and SO2 standards. In addition, substantial new information, based on observational data 

and rigorous atmospheric modeling, has become available regarding the role of both nitrogen and 

sulfur deposition in acidification of sensitive water bodies. This information is sufficient to 

inform the development of revised secondary standards for NOX and SOX to protect against the 

effects of acidification. While there is also new information available on the role of nitrogen 

deposition on nutrient enrichment effects in terrestrial and aquatic ecosystems, and the ISA 

concludes there is a causal relationship between NOX and nutrient enrichment effects, for this 

draft policy assessment, staff have focused on aquatic acidification effects due to the 

substantially greater amount of information available to inform the development of secondary 

standards for those effects. This is consistent with the available science and data, and also with 

the recommendations of CASAC, which stated that “EPA Staff is advised to focus on an AAPI 

standard driven by aquatic effects concerns.” There is not sufficient information at this time to 

develop secondary standards directly focused on protection of sensitive terrestrial ecosystems 

from acidification or sensitive terrestrial and aquatic ecosystems from adverse effects from 

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nutrient enrichment. Establishing standards that are multipollutant and ecologically relevant is 

an inherently complex process. We note that the aquatic acidification based standards are an 

important step in providing additional protections for sensitive ecosystems. While they do not 

provide complete protection for all sensitive ecosystems against all adverse effects, they will 

likely result in reductions in NOY and SO2 across broad regions of the U.S., resulting in 

decreased deposition and related effects for both terrestrial and aquatic ecosystems across all 

types of effects (see Chapter 6 for a broader discussion of this issue). 

Staff highlights the progress made in considering the joint nature of ecosystem responses 

to acidifying deposition of NOX and SOX, and notes that the ability to consider revisions to the 

NOX and SOX secondary standards has been enhanced by our ability to consider a joint standard 

for NOX and SOX to protect against aquatic acidification effects. The development of an 

appropriate form of the standard linked to a common indicator of aquatic acidification, ANC, is 

also a significant step forward, as it allows for development of a standard for aquatic 

acidification designed to provide generally the same degree of protection across the country, 

while still reflecting the underlying variability in ecosystem sensitivity to acidifying NOX and 

SOX deposition. 

 

9.1 CONSIDERATION OF ALTERNATIVE STANDARDS 

We begin by noting that the existing evidence continues to support existing NO2 and SO2

standards to protect against adverse effects associated with direct exposure of vegetation to gas 

phase NOX and SOX. 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 SOx and NOx. There is sufficient 

evidence to suggest that the levels of the current standards are likely adequate to protect against 

direct phytotoxic effects. As such, staff concludes that retaining the existing NO2 and SO2

standards to continue protection against these effects is appropriate. However, as discussed in 

Chapter 4, we also conclude that the existing secondary NOX and SOX standards are not adequate 

to provide protection of aquatic and terrestrial ecosystems against the effects of acidifying 

deposition. In response to this conclusion, staff considers a second overarching question: 

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What additional NOX and SOX standards are supported by the currently available scientific 

evidence and risk-based information, as reflected in the ISA and REA? 

To inform the answer to this overarching question, we have posed a series of more specific 

questions to aid in considering how the current NOX and SOX standards might be revised to 

provide requisite public welfare protection. 

Chapter 5 provided a conceptual framework for a secondary standard that is designed to 

provide protection of ecosystems against the effects associated with deposition of ambient 

concentrations of NOX and SOX. Chapter 5 also provided a discussion of potential options for the 

elements of a standard based on that conceptual framework, with a focus on the form of the 

standards. While we recognize the potential for significant impacts of current levels of NOX and 

SOX on terrestrial ecosystems and the effect of current levels of NOX on nutrient enrichment in 

sensitive aquatic ecosystems, we conclude that the currently available information is insufficient 

to develop either individual or joint standards to protect against these effects. We note that 

development of a standard for protection against terrestrial acidification may be appropriate 

using the same structure as we are proposing for aquatic acidification, using the Bc:Al ratio as 

the ecological indicator. However, the data needed to parameterize the form of such a standard 

is not sufficient at this time. As a result, we conclude that the current state of knowledge 

supports a standard to protect against the adverse effects associated with acidification of aquatic 

ecosystems. Such a standard is likely to provide some level of protection against other endpoints 

associated with deposition of N and S, but is not likely to adequately protect all sensitive 

terrestrial ecosystems or all N nutrient sensitive aquatic ecosystems. 

Building on the options discussed in Chapter 5, this section offers staff conclusions 

regarding the elements of the standard, including the indicators for NOX and SOX, the form of 

the standard, the averaging times, and presents for consideration options for target ANC levels 

associated with protection against specific ecological effects in aquatic ecosystems. Ultimately, 

the levels of AAPI considered by the Administrator should incorporate consideration of target 

levels of ANC, percent of waterbodies protected within defined spatial areas, and trajectories for 

ecosystem recovery, as well as uncertainties in the components of the AAPI. Associated with 

these elements of the AAPI are sets of NOy/(SO2+SO4) tradeoff curves which determine the 

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levels of ambient NOy and SO2+SO4 which will satisfy the level of the standard. The 

expression of these trade-off curves embodies the depositional load for a specified spatial area 

that is equivalent to the critical load for the waterbody in that area that represents a selected 

percentile (e.g. 95th percentile) of critical loads across waterbodies in the area, such that the 

selected percent of waterbodies in that area are expected to achieve an ANC at the target level. 

If a target load for a specific temporal period (e.g. by 2030) is evaluated instead of the critical 

load, then the depositional load represents the amount of deposition that is expected to achieve a 

target ANC value by a specific year for the selected percent of waterbodies in the area. The 

equivalent AAPI for an area can be calculated by inputting the values for each parameter of the 

AAPI equation for the selected percentile waterbody and the observed values of NOy and 

SO2+SO4. 

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These elements will be considered collectively in evaluating the protection from welfare 

effects associated with aquatic acidification afforded by alternative standards under 

consideration. In considering the currently available scientific and technical information, we 

consider both the information available in the last review and information that is newly available 

since the last review as assessed and presented in the ISA and RA prepared for this review (US 

EPA, 2008; US EPA, 2009). 

19 9.1.1. Indicators 

Staff concludes that the appropriate indicators for NOX and SOX, as described in detail in 

Chapter 5, are total NOy and the sum of SO2 and SO4, respectively. Total NOy includes all 

nitrogen oxides, including e.g. total reactive oxidized atmospheric nitrogen, defined as NOX (NO 

and NO2) and all oxidized NOX products: NOy = NO2 + NO + HNO3 + PAN +2N2O5 + 

HONO+ NO3 + organic nitrates + particulate NO3. The sum of SO2 and SO4 constitutes 

virtually the entire ambient air sulfur budget and SO2 and SO4 are measured routinely in 

existing monitoring networks. 

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As noted in Chapter 2 and Chapter 5, episodes of acidification and chronic acidification 

levels are associated with deposition over longer periods of time due to the storage and release of 

deposited nitrogen and sulfur in soils, snow, and ice. As a result, while episodic acidification 

may occur on much shorter timeframes, e.g. days to weeks, the cause of these episodes is largely 

due to shifts in hydrological flow paths, and the impact of these episodes is still determined by 

long term deposition of NOy and SO2+SO4 and associated long-term ANC, and thus the 

appropriate averaging time for ambient NOy and SO2+SO4 will be longer term. The averaging 

time for ambient concentrations of NOy and SO2+SO4 should be reflective of the long-term 

cumulative nature of deposition. CASAC supports using a three to five year averaging period 

“to help smooth out the year-to-year climatic variation in air concentration and deposition 

estimates.” (CASAC, 2010) 

13 9.1.3. Form 

The “form” of a standard defines the air quality statistic that is to be compared to the 

level of the standard in determining whether an area attains the standard. As discussed 

extensively in chapter 5, staff concludes that the current forms of the NOX and SOX secondary 

standards are not appropriate for addressing ecosystem acidification effects, and also concludes 

that a form that combines NOX and SOX levels with information on ecosystem sensitivity and 

nitrogen retention and uptake is most appropriate to maximize the likelihood of protecting 

sensitive ecosystems from the effects of acidification, without requiring standards that are more 

than requisite to provide that protection. Specifically, staff concludes that the Atmospheric 

Acidification Protection Index form as described in Chapter 5 is best suited to provide for 

protection against adverse effects due to acidifying deposition related to NOX and SOX. 

Within the AAPI, it is also appropriate to consider the specification of values of non-air 

quality parameters of the AAPI, including pre-industrial base cation weathering, nitrogen 

retention and uptake, runoff, and levels of reduced nitrogen deposition. As discussed in Chapter 

5, staff is proposing that the pre-industrial base cation weathering, nitrogen retention and uptake, 

and runoff values be determined by assessing the critical loads associated with a specific 

percentile of the waterbodies within defined spatial boundaries (as noted in Chapter 5, 

consideration is being given to a number of different methods for defining spatial boundaries). 

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The values of pre-industrial base cation weathering, nitrogen retention and uptake, and runoff 

values for a selected percentile waterbody are then used as the values in the form of the standard 

as realized for the specific ecoregion. 

The value of reduced nitrogen is initially set using deposition of NHx modeled using the 

CMAQ, evaluated for the period 2002-2005. Staff is considering the most appropriate spatial 

averaging extent for reduced nitrogen. Figure 9-1 shows spatially interpolated values of reduced 

nitrogen deposition based on 12km CMAQ modeling in the Eastern U.S. It is clear that in some 

locations, there is significant heterogeneity in reduced nitrogen deposition within ecoregion III 

boundaries. Given this information, two possible approaches to estimating reduced nitrogen 

values in the AAPI algorithm are 1) average reduced nitrogen deposition within an ecoregion, 

acknowledging that this will lead to uncertainties in the level of protection associated with levels 

of ambient NOY and SOX, or 2) allow for additional spatial refinement of sensitive areas to 

reflect the heterogeneity of reduced nitrogen deposition. This will result in multiple 

parameterizations of the AAPI within a single ecoregion. 

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Kriged NH3 Dep

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Figure 9-1. Spatially interpolated CMAQ estimates of deposition of reduced nitrogen 

(2002-2004 average) 

Unlike other parameters in the AAPI, reduced nitrogen is expected to change significantly over 

time because of the largely anthropogenic sources of reduced nitrogen deposition. In order to 

address this potential, we are exploring methods for specifying the standards in a way that would 

provide for updating the values of reduced nitrogen in the AAPI based on new modeling of 

reduced nitrogen deposition, following requirements for modeling established by EPA. 

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1 9.1.4. Considerations in Defining Options for the AAPI Standard 

Conclusions regarding an appropriate range of levels of the AAPI standard will be 

informed by considerations of the levels of the target ecological indicator ANC (related to levels 

of protection from effects of chronic and episodic acidification), relationships between target 

ANC levels and trajectories of recovery over time, and uncertainties in the various elements of 

the AAPI that affect the likelihood that the level of protection intended by a particular target 

ANC will be realized when atmospheric concentrations of NOy and SO2+SO4 fall below the 

tradeoff curve for that target ANC. Chapter 5 provides an assessment of the information related 

to selection of a target ANC, and the rationale for focusing additional considerations on specific 

target ANC levels, including 20 and 50 μeq/L. 

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The “levels” of the ambient indicators are determined by the selection of a level for the 

AAPI, as they represent the quantities of the ambient indicators that will result in the specified 

level of the AAPI. Those levels will vary for different locations depending on the nonatmospheric related characteristics of the ecosystem, the level of reduced nitrogen in the 

ecosystem, and the atmospheric transformation ratios (TNOy and T(SO2+SO4)). 

The secondary NAAQS will reflect the public welfare policy judgments of the 

Administrator, based on the science, as to the level of air quality which is requisite to protect the 

public welfare from any known or anticipated adverse effects associated with the pollutant in the 

ambient air. The exposure and risk assessment provide information regarding the effects 

associated with a number of different welfare endpoints at different levels of air quality, 

expressed in terms of the joint multiyear mean concentrations of NOy and SO2+SO4 determined 

such that specific levels of ecosystem protection (for example, ANC greater than 50 μeq/L) are 

met. Staff also recognizes that in certain naturally acidic ecosystems, even though the ecological 

benchmarks are exceeded, e.g. ANC may be quite low; NOy and SO2+SO4 are not contributing 

to effects because those systems have chronic natural acidity and will not benefit from reductions 

in atmospheric deposition. The secondary NAAQS are not intended to provide protection in 

these types of naturally acidic systems. As a result, in our determination of appropriate lakes and 

streams to include in the populations of critical loads that determine protective NOy and 

SO2+SO4 levels, we apply filters to remove lakes and streams that are naturally acidic or 

acidified due to mine drainage. The secondary NAAQS are focused on providing protection in 

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areas where ambient NOy and SO2+SO4 are resulting in effects in ecosystems with low natural 

levels of acidification that are highly sensitive to additional inputs of acid deposition. 

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Staff concludes that ecosystem effects of NOy and SO2+SO4 deposition in aquatic 

ecosystems are an important public welfare effect of concern, based on the types and extent of 

ecosystem services likely to be affected by deposition, as well as the location of some aquatic 

ecosystems within state and national protected lands, including Class I national parks and 

wilderness areas (Chapter 3). 

9.1.4.1 Target ANC Level 

In reaching staff conclusions regarding target ANC levels that are appropriate to consider for an 

AAPI-based standard, staff take into account the currently available scientific information 

including: evidence from field and laboratory studies, including evidence of effects in highly 

sensitive ecosystems and estimates of risk reductions associated with alternative annual standard 

levels, as well as the related limitations and uncertainties associated with this information as 

presented and discussed more fully in the ISA and RA (US EPA, 2008; US EPA, 2009). In 

developing conclusions regarding the target ANC, we evaluated both evidence and risk based 

information. In addition, we consider information on target ANC and pH levels used by other 

organizations that have established critical loads for protection of aquatic ecosystems from 

effects of acidification. 

We conclude that it is appropriate to define NOX and SOX standards that will provide 

generally consistent protection for acid sensitive lakes and streams across the country. In order 

to do so, we focus attention on considering target ANC levels that will lead to a level of AAPI 

that provides the same protection for sensitive aquatic ecosystems throughout the U.S. The 

result of focusing on a nationally protective level of AAPI is a set of varying NOy/(SO2+SO4) 

tradeoff curves across the U.S. reflecting that NOy and SO2+SO4 affect acidification in different 

ways depending on underlying ecosystem characteristics and levels of reduced nitrogen, such 

that the same AAPI is calculated with differing levels of NOy and SO2+SO4. This approach 

recognizes that changes in air quality to meet the standards may reflect differing combinations of 

NOy and SO2+SO4 leading to the same level of the AAPI. 

Based on our analyses of risks of impacts on aquatic species diversity and fitness and on 

the basis of the scientific effects literature, we conclude that achieving a target ANC of 50 μeq/L 

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would substantially decrease the effects of acidification due to NOy and SO2+SO4 on aquatic 

ecosystems, decreasing the risk of losses in biodiversity and mortality in fish and other aquatic 

organisms, and improving the overall health of aquatic ecosystems. Additionally, it is 

anticipated that achieving a target ANC of 50 μeq/L would provide increased protection from 

NOy and SO2+SO4 in areas with higher levels of variability in ecosystem sensitivity due to 

variability in meteorology, bedrock geology, topography, land use characteristics, or reduced 

nitrogen deposition. 

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It is recognized, however, that a achieving a target ANC of 50 μeq/L would likely not 

protect the most sensitive aquatic ecosystems or species within those ecosystems from the effects 

of NOy and SO2+SO4. At ANC levels below 100 μeq/L, while overall health of an aquatic 

community can be maintained, ANC levels are expected to be such that fish fitness and 

community diversity begin to decline. At ANC levels between 100 and 50 μeq/L, ANC levels 

are expected to be such that the fitness of sensitive species (e.g., brook trout, zooplankton) also 

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9.1.4.2 Target Percent of Waterbodies to Meet a Target ANC Level 

The appropriate range of levels of AAPI is also informed by the selection of a target 

percent of protection for waterbodies within particular acid sensitivity classes or ecoregions (see 

Chapter 5). More specifically, the greater percentage of waterbodies that are to be protected 

generally indicates a greater likelihood that sensitive waterbodies will achieve a target ANC 

level, and as such, in setting the level of the standard, there will be less need to reflect 

uncertainty in the likelihood that those sensitive waterbodies will be protected. 

 It is also important to consider that while the target ANC level will not be met in all 

sensitive waterbodies if a target percent less than 100 is selected, all waterbodies will realize 

some level of protection due to decreases in NOy and SO2+SO4 to meet the target ANC in 

targeted waterbodies. 

 Additional analyses of the implications of alternative target percentages of waterbodies 

are underway and are expected to be completed for inclusion in the final PA. 

9.1.4.2 Additional Considerations Related to Developing Options for the AAPI Form 

of the Standard 

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In developing options for the standard, consideration should of the degree to which any 

specific AAPI would lead to achieving the desired ANC level, and a judgment as to the degree of 

protection of public welfare that is warranted. These considerations should incorporate a wide 

number of factors, including the percent of water bodies within acid sensitive areas that the 

Administrator determines should be protected at the targeted ANC level, as well as consideration 

of achieving desired levels of protection within generational timeframes, e.g., 20 to 40 years, 

concerns about protection against episodic acidification events, and uncertainties in the modeling 

of critical loads, nitrogen uptake and retention, reduced nitrogen deposition, and relationships of 

atmospheric concentrations to deposition. In considering options for the standard that would 

reflect consideration for providing requisite welfare protection against know or anticipated 

effects, we believe that while the available information is insufficient to set separate standards 

for terrestrial acidification and aquatic and terrestrial nutrient enrichment effects, it is also 

appropriate to consider the evidence of those effects and the likelihood for co-protection 

provided by standards targeted at protection against effects of aquatic acidification (Chapter 6). 

Chapter 7 provided a summary and synthesis of critical uncertainties and implications for 

the standards. While many uncertainties cannot be quantitatively assessed, and as such cannot be 

used to recommend specific quantitative changes to the AAPI, there are several uncertainty 

analyses which give some insight into the magnitude and direction of the uncertainty. For 

example, uncertainty analysis of the MAGIC model of critical loads indicates that modeled preindustrial ANC (which informs the distribution of critical loads on which NOy and SO2+SO4 

levels are based by establishing the natural ability of an aquatic ecosystem to neutralize acid 

inputs) has 95 percent confidence intervals that are 10 percent higher (or lower) than the mean 

estimate for lakes, and 5 percent higher (or lower) than the mean estimate for streams. Similar 

uncertainty exists regarding overall uncertainty in the models used to generate critical loads for 

determining the NOy and SO2+SO4 tradeoff curves. 

The deposition of reduced nitrogen is a critical input to the AAPI and has a large 

expected uncertainty due to the use of CMAQ modeling which relies on uncertain chemistry and 

uncertain inventories of ammonia emissions. Much of the uncertainty introduced by reduced 

nitrogen deposition can be decreased by improvements in measurements of reduced nitrogen 

deposition and improvements in the emissions inventories of ammonia and characterization of 

ammonia chemistry within the CMAQ modeling. These improvements are underway, and our 

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form of the standard is designed to allow for dynamic updates to reduced nitrogen deposition 

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5 9.1.5 Additional protections for ecosystems against the effects of terrestrial acidification 

 and terrestrial and aquatic nutrient enrichment 

While we are not basing the elements of this standard primarily on consideration of 

effects other than aquatic acidification, our approach recognizes that some level of protection 

against effects of acidification in terrestrial ecosystems and effects of nutrient enrichment in 

terrestrial and aquatic ecosystems is likely to be realized through changes in air quality to meet 

the AAPI standard. We recognize that an annual standard focused on aquatic acidification 

cannot be expected to offer protection against all of the welfare effects from NOy and SO2+SO4, 

especially in areas that are sensitive to nutrient enrichment but are not acid sensitive. However, 

based on the information available in the ISA and REA, we conclude that the available 

information is not sufficient to set a complementary standard to provide protection against 

additional effects of NOy through nutrient enrichment, and that additional research is necessary 

to support the setting of such a standard, especially in the areas of identifying the specific 

impacts of decreases in atmospheric nitrogen deposition (see chapter 7). CASAC has noted that 

our current framework, with appropriate modifications, should be applicable to developing 

standards to provide protection against acidification effects in sensitive terrestrial ecosystems. 

However, staff has concluded that the current data is not sufficient to develop a separate 

terrestrial acidification based standard at this time. A primary limitation is the identification of 

specific levels of harm associated with the BC/Al ecological indicator. 

Sensitive terrestrial ecosystems that are located in watersheds with acid sensitive water 

bodies are likely to receive the most protection under an aquatic acidification targeted standard. 

Terrestrial ecosystems outside of these watersheds are likely to see some level of protection, but 

will not realize targeted changes in ambient NOy and SO2+SO4 (Chapter 6). 

Chapter 6 evaluated the relative protection for terrestrial ecosystems in areas from 

meeting a target ANC of 50 μeq/L compared to meeting a target Bc:Al ratio of 10. Critical 

loads for N and S were compared to determine which of the targets would result in a more 

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stringent critical load. Over half of the watersheds had a lower critical load to meet the ANC 

target compared to the critical load to meet the Bc:Al ratio target. As a result, those watersheds 

are likely to be protected from both terrestrial and aquatic impacts when the ANC target is met. 

When the water bodies are more sensitive to deposition (“highly sensitive” or “moderately 

sensitive”), the aquatic critical acid loads generally provide a greater level of protection against 

acidifying nitrogen and sulfur deposition in the watershed. 

The tradeoff curves for NOy and SO2+SO4 that are associated with protection against 

aquatic acidification also provide bounding conditions for nitrogen that can be compared against 

benchmarks of effects associated with nitrogen deposition in sensitive terrestrial and aquatic 

ecosystems. Achieving an ANC level of 50 μeq/L for 90 percent of lakes and streams 

nationwide would provide some protection against leaching in northeast forests, but would need 

to be lower to protect California coastal sage scrub, lichens in mixed conifer forests, alpine lake 

communities, and Minnesota grasslands (Section 6.2). In the case of aquatic nutrient 

enrichment, comparison of maximum allowable NOX levels on the tradeoff curves with 

deposition requirements to meet the Chesapeake Bay TMDL shows that without further 

restrictions on NOy concentrations, standards set to protect against aquatic acidification will not 

be protective against effects of aquatic nutrient enrichment in the Chesapeake Bay. 

9.1.6 Summary of options 

 To facilitate evaluation of the elements of the standard and staff conclusions regarding 

those elements, we have constructed a summary table showing the elements of the AAPI, options 

for each element, and staff conclusions where appropriate. 

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[Table 9-1 to be provided] 

9.2 CONCLUSIONS 

The following secondary NAAQS conclusions encompass the breadth of policy-relevant 

considerations described in this policy assessment. We note that staff conclusions to be 

presented in the final PA will consider input received from CASAC and the public on this second 

draft PA. We recognize that selecting from among alternative standards will necessarily reflect 

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consideration of the qualitative and quantitative uncertainties inherent in the relevant evidence 

and in the assumptions that underlie the quantitative risk assessment. In identifying these 

alternative secondary standards and ranges of levels for consideration, we are mindful that the 

Clean Air Act requires standards to be set that are requisite to protect public from known or 

anticipated adverse effects, such that the standards are to be neither more nor less stringent than 

necessary. Thus, the Act does not require that the NAAQS be set at no effect levels, but rather at 

levels that avoid adverse effects on public welfare: 

(1) Based on the policy-relevant findings from the ISA described in Chapter 2, and while 

recognizing that important uncertainties and research questions remain, staff conclude that great 

progress has been made since the last reviews of the secondary standards for NOX and SOX. We 

generally find support in the available effects-based evidence for consideration of NOx and SOx 

standards that are at least as protective as the current standard and do not find support for 

consideration of NOX and SOX standards that are less protective than the current standard. The 

staff also concludes that consideration of joint standards for NOX and SOX is appropriate given 

the common atmospheric processes governing the deposition of NOX and SOX to sensitive 

ecosystems, and given the combined effects of N and S deposition on acidification of soil and 

water. 

 (2) Staff concludes that ambient NOX is a significant component of atmospheric nitrogen 

deposition, even in areas with relatively high rates of deposition of reduced nitrogen. Staff make 

this conclusion based on the analysis in Chapter 3 of the REA, which provides a thorough 

assessment of the contribution of NOX to nitrogen deposition throughout the U.S., and the 

relative contributions of ambient NOX and reduced forms of nitrogen. 

(3) Staff concludes based on the case study results provided in the REA, that current 

levels of NOX and SOX are associated with deposition that leads to ANC values below 

benchmark values that cause ecological harm and losses in ecosystem services. Staff concludes 

that the evidence and risk assessment support strongly a relationship between atmospheric 

deposition of NOX and SOX and ANC, and that ANC is an excellent indicator of aquatic 

acidification. Staff also concludes that at levels of deposition associated with NOX and SOX

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, which is 

associated with reductions in recreational fishing services. Although there are many other 

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ecosystem services potentially affected by reductions in ANC, including subsistence fishing, 

natural habitat provision, and biological control, confidence in the specific translation of ANC 

values to these additional ecosystem services is much lower. 

(4) Losses in aquatic resources associated with ANC levels below 50 μeq/L are clearly 

associated with significant losses in economic value. Based on the best available data, just in 

New York., increasing ANC levels to 50 in the Adirondacks is estimated to result in $300 to 

$800 million in annual benefits in 2006 dollars. This estimate represents only a fraction of the 

total economic value of ecosystem damages as many impacted resources are not amenable to 

economic valuation methods. In addition, economic damages are also likely to occur in other 

areas affected by acidification, including New England, the Appalachian Mountains (northern 

Appalachian Plateau and Ridge/Blue Ridge region), and the Upper Midwest. Staff concludes 

that reducing acidifying deposition of NOX and SOX will result in improvements in public 

welfare by increasing the quantity and quality of ecosystem services, including recreational 

fishing and other services associated with improved water quality. 

(5) Staff initially concludes based on the case study results that current levels of ambient 

NOx and SOx are associated with deposition that leads to Bc:Al values below benchmark values 

that cause ecological harm and losses in ecosystem services. Staff concludes that the evidence 

and risk assessment support strongly a relationship between atmospheric deposition of NOX and 

SOX and Bc:Al, and that Bc:Al is a good indicator of terrestrial acidification. Staff also 

concludes that at levels of deposition associated with NOX and SOX concentrations at or below 

the current standards, Bc:Al levels are expected to be below benchmark values that are 

associated with significant losses in tree health and growth, which are associated with reductions 

in timber production. While there are many other ecosystem services, including maple syrup 

production, natural habitat provision, and regulation of water, climate, and erosion, potentially 

affected by reductions in Bc:Al, confidence in the specific translation of Bc:Al values to these 

additional ecosystem services is much lower. 

(6) On the basis of the acidification and nutrient enrichment effects that have been 

observed to still occur under current ambient conditions and those predicted to occur under the 

scenario of just meeting the current secondary NAAQS, staff concludes that the current 

secondary NAAQS are inadequate to protect the public welfare from known and anticipated 

adverse welfare effects from aquatic and terrestrial acidification associated with deposition of 

September, 2010 Draft –Do Not Quote or Cite 9-16
NOX and SOX.. As discussed above, this conclusion derives from several lines of evidence. Staff 

also concludes that the current NOX and SOX secondary standards are adequate to protect against 

direct gas-phase effects on vegetation, and as such, should be retained to preserve protection 

against these welfare effects.

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(7) Staff has concluded, based on the completeness of the available evidence and 

quantitative risk information, that effects due to aquatic acidification are most suitable for 

defining additional secondary standards for NOX and SOX. Staff notes that in developing a 

standard designed to protect against the effects of aquatic acidification due to deposition of NOX

and SOX, the resulting standards may not provide protection against known effects associated 

with terrestrial acidification and with nutrient enrichment in sensitive aquatic and terrestrial 

ecosystems. 

 (8) It is appropriate to consider using indicators other than NO2 and SO2 as the indicators 

for an additional standard that is intended to address the ecological effects associated with 

deposition of NOX and SOX to sensitive ecosystems. Given the reasons discussed in Chapters 2, 

4, and 5 of this policy assessment, staff concludes that NOX (oxides of nitrogen, the definition in 

section 302 (v) of the CAA), is best represented by the atmospheric indicator NOy, defined as 

NO2 + NO + HNO3 + PAN +2N2O5 + HONO+ NO3 + organic nitrates + particulate NO3 is the 

more appropriate indicator of oxides of nitrogen, and that SO2+SO4 is the more appropriate 

indicator of oxides of sulfur for purposes of a secondary standard addressing aquatic 

acidification. 

(9) It is appropriate to use multi-year averages of concentrations of NOy and SO2+SO4 as 

the averaging times for the secondary standards, based on the chronic nature of acidification, and 

the protection against episodic acidification provided by a standard based on annual average 

concentrations. Averaging periods in the range of 3 to 5 years are most appropriate. 

(10) It is appropriate to consider adding a new standard with a different form for NOy

and SO2+SO4 as the current form does not take into account the linkages between NOx and SOx 

in the causation of effects associated with acidification of aquatic ecosystems. Based on the 

causal linkages between NOX and SOX, deposition of N and S, and the indicator of acidification, 

ANC, staff concludes that a standard with a form specified as an atmospheric acidification 

protection index (AAPI) should be added. A standard with this form would reflect the important 

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roles of underlying ecosystem characteristics, determinants of deposition, and deposition of 

reduced nitrogen in determining the potential effects from deposition of NOX and SOX. 

(11) Staff has concluded, based on the evidence and risk based information, and 

consideration of information related to definitions of adversity, that 

a) a target level of ANC of 20 μeq/L will protect against significant losses in 

fish mortality in many sensitive lakes, but will place less weight on protection against 

losses in aquatic biodiversity, and will be less protective against potential acidification 

episodes, 

b) a target level of ANC of 50 μeq/L will protect against significant mortality 

in aquatic organisms and loss of fish health and biodiversity in sensitive lakes and 

streams, and will give weight to considerations of uncertainties in the time to recovery of 

aquatic ecosystems, 

c) target levels of ANC above 50 μeq/L may provide additional protection 

against declines in fitness of sensitive species (e.g., brook trout, zooplankton), however, 

overall health of aquatic communities may not be impacted. 

September, 2010 Draft –Do Not Quote or Cite 9-18
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9.3 REFERENCES 

National Research Council. 2004. Air Quality Management in the United States. National 

Academies Press, Washington, D.C. 

CASAC (2010). Review of the Policy Assessment for the Review of the Secondary National 

Ambient Air Quality Standards for NOx and SOx: First Draft (March 2010). EPACASAC-10-014. June 22. 

US EPA (2008) U.S. EPA. Integrated Science Assessment (ISA) for Oxides of Nitrogen and 

Sulfur Ecological Criteria (Final Report). U.S. Environmental Protection Agency, 

Washington, D.C., EPA/600/R-08/082F 

US EPA (2009) Risk and Exposure Assessment for Review of the Secondary National Ambient 

Air Quality Standards for Oxides of Nitrogen and Oxides of Sulfur-Main Content - Final 

Report. U.S. Environmental Protection Agency, Washington, D.C., EPA-452/R-09-008a 
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ADDITIONAL INFORMATION FOR CHAPTER 5 

A.1Conceptual Design of the Standard 

This is supplemental information to support the discussion of the conceptual 

design of the standard that is presented in Chapter 5 of the Policy Assessment Document. 

The aquatic acidification analyses developed in the REA used a number of different 

models and calculation techniques that are important for the development of the standard. 

The goal of this Appendix is to summarize information from the REA analysis that is 

most relevant to the Policy Assessment. A brief summary of the REA analyses are 

presented in section A.1.1. In section A.1.2 there is a general summary and technical 

discussion of the critical loads modeling approaches that were used in the REA, followed 

by a brief description of MAGIC model data requirements. 

A.1.1 Technical summary of methods used in the REA Aquatic Acidification 

analysis 

The aquatic acidification analysis is presented in Chapter 4 and Appendix 4 of 

the REA. The analysis uses multiple techniques to show the relationship between ANC 

and NOx and SOx deposition, as well as determine the current level of risk to water 

bodies that occur in sensitive areas. A brief summary of the techniques and objectives of 

the REA analysis is given in Table A-1. 

Table A-1. Brief summary of objects and methods used in the REA Aquatic Acidification 

analysis. 

Technique Objectives 

Time-series 

graphs of 

current 

conditions 

1 Data from monitoring networks collected from 1990 to 2006 were 

plotted to show trends in concentrations of pollutants, deposition and 

acidification for each case study site. The data included surface water 

concentration of nitrate, sulfate and ANC; deposition of sulfate and 

nitrate; as well as air concentration of SOx, NOx and NH4 

MAGIC 1 Used to estimate the relationship between ANC values and 

anthropogenic NOx and SOx emission from the past (preacidification 

~1860), present (2002 and 2006) and projected into the future (2020 

and 2050). Analysis included 44 lakes from Adirondacks and 60 

streams from Shenandoah. 

2 Used to develop input parameters for critical loads modeling (i.e. 

weathering rates) 

3 Used for uncertainty analysis 

Critical Loads 

modeling 

1 SSWC and FAB models used to calculate critical loads for critical 

limits of ANC = 0, 20, 50, 100 

September, 2010 Draft – Do Not Quote or Cite A-1
Table A-1. Brief summary of objects and methods used in the REA Aquatic Acidification 

analysis. 

2 Critical loads for ANC critical limits calculated for 169 lakes in the 

Adirondacks and 60 streams in the Shenandoah using water quality 

data from monitoring sites collected in 2006 

3 Critical loads exceedences calculated by comparing the critical loads 

that were calculated by SSWC with deposition data from NADP for 

wet deposition and CMAQ for dry deposition, both for the year 2002 

Regional 

Extrapolation 

1 117 of the critical loads calculated for the Adirondacks were 

extrapolated to lakes defined by the New England EMAP probability 

survey, representing 1842 lakes, to infer the # of lakes that exceeded 

their critical load 

2 69 of the critical loads calculated for the Shenandoah were 

extrapolated to 330 streams based on bed rock geology classification. 

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A.1.2 Technical summary of critical loads modeling in the REA 

The critical load of acidity for lakes or streams was derived from present-day 

water chemistry using a combination of steady-state models. Both the Steady-State Water 

Chemistry (SSWC) model and First-order Acidity Balance model (FAB) is based on the 

principle that excess base-cation production within a catchment area should be equal to or 

greater than the acid anion input, thereby maintaining the ANC above a preselected level 

(Reynolds and Norris, 2001; Posch et al. 1997). These models assume steady-state 

conditions and assume that all SO4

2– in runoff originates from sea salt spray and 

anthropogenic deposition. Given a critical ANC protection level, the critical load of 

acidity is simply the input flux of acid anions from atmospheric deposition (i.e., natural 

and anthropogenic) subtracted from the natural (i.e., preindustrial) inputs of base cations 

in the surface water. Final Risk and Exposure Assessment September 2009 Appendix 4, 

Attachment A – 15 Aquatic Acidification Case Study Atmospheric deposition of NOx and 

SOx contributes to acidification in aquatic ecosystems through the input of acid anions, 

such as NO3- and SO4

2– The acid balance of headwater lakes and streams is controlled by 

the level of this acidifying deposition of NO3- and SO4

2– and a series of biogeochemical 

processes that produce and consume acidity in watersheds. The biotic integrity of 

freshwater ecosystems is then a function of the acid-base balance, and the resulting 

acidity-related stress on the biota that occupy the water. The calculated ANC of the 

surface waters is a measure of the acid-base balance: 

ANC = [BC]* - [AN]* (1) 

September, 2010 Draft – Do Not Quote or Cite A-2
where [BC]* and [AN]* are the sum of base cations and acid anions (NO3- and SO4

2–), 

respectively. Equation (1) forms the basis of the linkage between deposition and surface 

water acidic condition and the modeling approach used. Given some “target” ANC 

concentration [ANClimit]) that protects biological integrity, the amount of deposition of 

acid anions (AN) or depositional load of acidity CL(A) is simply the input flux of acid 

anions from atmospheric deposition that result in a surface water ANC concentration 

equal to the [ANClimit] when balanced by the sustainable flux of base cations input and 

the sinks of nitrogen and sulfur in the lake and watershed catchment. 

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Critical loads for nitrogen and sulfur (CL(N) + CL(S) ) or critical load of acidity 

CL(A) were calculated for each waterbody from the principle that the acid load should 

not exceed the nonmarine, nonanthropogenic base cation input and sources and sinks in 

the catchment minus a neutralizing to protect selected biota from being damaged: 

CL(N) + CL(S) or CL(A) = BC*dep + BCw – Bcu – AN - ANClimit (2) 

Where, 

 BC*dep = (BC*=Ca*+Mg*+K*+Na*), nonanthropogenic deposition flux of base cations 

BCw = the average weathering flux, producing base cations 

Bcu (Bc=Ca*+Mg*+K*) = the net long-term average uptake flux of base cations in the 

biomass (i.e., the annual average removal of base cations due to harvesting) 

AN = the net long-term average uptake, denitrification, and immobilization of nitrogen 

anions (e.g. NO3-) and uptake of SO4

2– 

ANClimit = the lowest ANC-flux that protects the biological communities. 

Since the average flux of base cations weathered in a catchment and reaching the 

lake or streams is difficult to measure or compute from available information, the average 

flux of base cations and the resulting critical load estimation were derived from water 

quality data (Henriksen and Posch, 2001; Henriksen et al., 1992; Sverdrup et al., 1990). 

Weighted annual mean water chemistry values were used to estimate average base cation 

fluxes, which were calculated from water chemistry data collected from the Temporally 

Integrated Monitoring of Ecosystems (TIME)/Long-Term Monitoring (LTM) monitoring 

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networks, that include Adirondack Longterm Monitoring (ALTM), Virginia Trout Stream 

Sensitivity Study (VTSSS), and the Shenandoah Watershed Study (SWAS), and 

Environmental Monitoring and Assessment Program (EMAP) (see Section 4.1.2.1 of 

Chapter 4). 

The preacidification nonmarine flux of base cations for each lake or stream, BC*0, is 

BC*0 = BC*dep + BCw - Bcu (3) 

Thus, critical load for acidity can be rewritten as 

CL(N) + CL(S) = BC*0 – AN - ANClimit = Q.([BC*]0 – [AN] - [ANC]limit), (4) 

where the second identity expresses the critical load for acidity in terms of catchment 

runoff (Q) m/yr and concentration ([x] = X/Q). The sink of nitrogen in the watershed is 

equal to the uptake (Nupt), immobilization (Nimm), and denitrification (Nden) of 

nitrogen in the catchment. Thus, critical load for acidity can be rewritten as 

CL(N) + CL(S) = {fNupt + (1 − r)(Nimm + Nden)} + ( [BC]0* − [ANClimit])Q (5) 

where f and r are dimensionless parameters that define the fraction of forest cover in the 

catchment and the lake/catchment ratio. The in-lake retention of nitrogen and sulfur was 

assumed to be negligible. Equation 5 described the FAB model that was applied when 

sufficient data was available to estimate the uptake, immobilization, and denitrification of 

nitrogen and the neutralization of acid anions (e.g. NO3-) in the catchment. In the case 

were data was not available, the contribution of nitrogen anions to acidification was 

assumed to be equal to the nitrogen leaching rate (Nleach) into the surface water. The 

flux of acid anions in the surface water is assumed to represent the amount of nitrogen 

that is not retained by the catchment, which is determined from the sum of measured 

concentration of NO3- and ammonia in the stream chemistry. This case describes the

SSWC model and the critical load for acidity is 

CL(A) = Q.([BC*]0 – [ANC]limit) (6) 

where the contribution of acid anions is considered as part of the exceedances calculation 

(see Section 1.2.5, below). For the assessment of current condition in both case study 

September, 2010 Draft – Do Not Quote or Cite A-4
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areas, the critical load calculation described in Equation 6 was used for most lakes and 

streams. The lack of sufficient data for quantifying nitrogen denitrification and 

immobilization prohibited the wide use of the FAB model. In addition, given the 

uncertainty in quantifying nitrogen denitrification and immobilization, the flux of 

nitrogen anions in the surface water was assumed to more accurately reflect the 

contribution of NO3- to acidification. Several major assumptions are made: (1) steadystate conditions exist, (2) the effect of nutrient cycling between plants and soil is 

negligible, (3) there are no significant nitrogen inputs from sources other than 

atmospheric deposition, (4) ammonium leaching is negligible because any inputs are 

either taken up by biota or adsorbed onto soils or nitrate compounds, and (5) longterm 

sinks of sulfate in the catchment soils are negligible. 

A.1.2.1 Preindustrial Base Cation Concentration 

Present-day surface water concentrations of base cations are elevated above their 

steadystate preindustrial concentrations because of base cation leaching through ion 

exchange in the soil due to anthropogenic inputs of SO4

2– to the watershed. For this 

reason, present-day surface water base cation concentrations are higher than natural or 

preindustrial levels, which, if not corrected for, would result in critical load values not in 

steady-state condition. To estimate the preacidification flux of base cations, the present 

flux of base cations was estimated, 

BC*t, given by BC*t = BC*dep + BCw – Bcu +BCexc, (7) 

Where BCexc = the release of base cations due to ion-exchange processes. Assuming 

that deposition, weathering rate, and net uptake have not changed over time, BCexc can 

be obtained by subtracting Equation 5 from Equation 7: 

BCexc = BC*t – BC*0 (8) 

This present-day excess production of base cations in the catchment was related to the 

long-term changes in inputs of nonmarine acid anions (ΔSO*2 + ΔNO3) by the F-factor 

(see below): 

September, 2010 Draft – Do Not Quote or Cite A-5
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BCexc = F (ΔSO*2 + ΔNO3) (9) 

For the preacidification base cation flux, solving Equation 5 for BC*0 and then 

substituting Equation 8 for BCexc and explicitly describing the long-term changes in 

nonmarine acid ion inputs: 

BC*0 = BC*t – F (SO*4,t - SO*4,0 + NO*3,t - NO*3,0) (10) 

The preacidification NO3- concentration, NO*3,0, was assumed to be zero. 

A.1.2.2 F-factor 

An F-factor was used to correct the concentrations and estimate preindustrial base 

concentrations for lakes in the Adirondack Case Study Area. In the case of streams in the 

Shenandoah Case Study Area, the preindustrial base concentrations were derived from 

the MAGIC model as the base cation supply in 1860 (hindcast) because the F-factor 

approach is untested in this region. An F-factor is a ratio of the change in nonmarine base 

cation concentration due to changes in strong acid anion concentrations (Henriksen, 

1984; Brakke et al., 1990): 

F =([BC*]t - [BC*]0)/([SO4*]t - [SO4*]0 + [NO3*]t - [NO3*]0), (12) 

where the subscripts t and 0 refer to present and preacidification conditions, respectively. 

If F=1, all incoming protons are neutralized in the catchment (only soil acidification); at 

F=0, none of the incoming protons are neutralized in the catchment (only water 

acidification). The F-factor was estimated empirically to be in the range 0.2 to 0.4, based 

on the analysis of historical data from Norway, Sweden, the United States, and Canada 

(Henriksen, 1984). Brakke et al. (1990) later suggested that the F-factor should be a 

function of the base cation concentration: 

F = sin (π/2 Q[BC*]t/[S]) (13) 

September, 2010 Draft – Do Not Quote or Cite A-6
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where 

Q = the annual runoff (m/yr). [S] = the base cation concentration at which F=1; and for 

[BC*]t>[S] F is set to 1. For Norway [S] has been set to 400 milliequivalents per cubic 

meter (meq/m3)(circa.8 mg Ca/L) (Brakke et al., 1990). The preacidification SO42- 

concentration in lakes, [SO4*]0, is assumed to consist of a constant atmospheric 

contribution and a geologic contribution proportional to the concentration of base cations 

(Brakke et al., 1989). The preacidification SO42- concentration in lakes, [SO4*]0 was 

estimated from the relationship between [SO42-]o* and [BC]t* based on work completed 

by Henriksen et al., 2002 as described by the following equation: 

[SO42-]o* = 15 + 0.16 * [BC]t* (14) 

September, 2010 Draft – Do Not Quote or Cite A-7
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Table A-2 Illustrates SSWC Approach – Environmental Variables 

CL(A) = BC*

dep + BCw – Bcu – ANClimit 

CL(A) = Q.([BC*]0 – [ANC]limit)

Variable 

Code 

Description Source 

1 BC*

dep Sum (Ca*+Mg*+K*+Na*), nonanthropogenic 

deposition flux of base cations 

Wet NADP and Dry 

CASTNET 

2 BCw Average weathering flux of base cations Calculated (5-17) 

3 Bcu Sum (Ca+Mg+K), the net long-term average 

uptake flux of base cations in the biomass 

USFS-FIA data 

4 ANClimit Lowest ANC-flux that protects the biological 

communities 

Set 

5 Ca* Sea Salt corrected Surface water concentration 

(μeq/L) growing season average. (Ca – (CL × 

0.0213)) 

Water quality data 

6 Mg* Sea Salt corrected Surface water concentration 

(μeq/L) growing season average. (Mg – (CL 

× 0.0669)) 

Water quality data 

7 Na* Sea Salt corrected Surface water concentration 

(μeq/L) growing season average. (Na – (CL × 

0.557)) 

Water quality data 

8 K* Sea Salt corrected Surface water concentration 

(μeq/L) growing season average. (K – (CL × 

0.0.0206)) 

Water quality data 

9 SO4

* Sea Salt corrected Surface water concentration 

(μeq/L) growing season average. (SO4 – (CL × 

0.14)) 

Water quality data 

10 CL Surface water concentration (μeq/L) growing 

season average. 

Water quality data 

11 SO4

* Surface water concentration (μeq/L) growing 

season average. 

Water quality data 

12 NO3

* Surface water concentration (μeq/L) growing 

season average. 

Water quality data 

13 Q The annual runoff (m/yr) USGS 

14 [BC*]0 Preindustrial flux of base cations in surface 

water, corrected for sea salts 

Calculated from water 

quality data 

15 [SO4

*

]0 Preindustrial flux of sulfate in surface water, 

corrected for sea salts 

Estimated 

16 [NO3

*

]0 Preindustrial flux of nitrate, corrected for sea 

salts 

Equal to 0 

17 F Calculated factor Fix values 

September, 2010 Draft – Do Not Quote or Cite A-8
Table A-2 FAB Approach – Environmental Variables 

DL(N) + DL(S) = {fNupt + (1 − r)(Nimm + Nden) + (Nret + Sret)} + ( [BC]0

* − [ANClimit])Q

Variable 

Code 

Description Source 

1 Ndepo Total N deposition NADP/CMAQ 

2 ANClimit Lowest ANC-flux that protects the biological 

communities 

Set 

3 [BC*]0 Preindustrial flux of base cations in surface 

water, corrected for sea salt 

Calculated from 

water quality data 

4 Ca* Sea Salt corrected Surface water concentration 

(μeq/L) growing season average. (Ca – (CL × 

0.0213)) 

Water quality data 

5 Mg* Sea Salt corrected Surface water concentration 

(μeq/L) growing season average. (Mg – (CL × 

0.0669)) 

Water quality data 

6 Na* Sea Salt corrected Surface water concentration 

(μeq/L) growing season average. (Na – (CL × 

0.557)) 

Water quality data 

7 K* Sea Salt corrected Surface water concentration 

(μeq/L) growing season average. (K – (CL × 

0.0.0206)) 

Water quality data 

8 SO4

* Sea Salt corrected Surface water concentration 

(μeq/L) growing season average. (SO4 – (CL × 

0.14)) 

Water quality data 

9 CL Surface water concentration (μeq/L) growing 

season average. 

Water quality data 

10 SO4

* Surface water concentration (μeq/L) growing 

season average. 

Water quality data 

11 NO3

* Surface water concentration (μeq/L) growing 

season average. 

Water quality data 

12 Q The annual runoff (m/yr) USGS 

13 f f is a dimensionless parameter that define the 

fraction of forest cover in the catchment 

14 r r is a dimensionless parameter that define the 

lake/catchment ratio 

14 Nret The in-lake retention of nitrogen Estimated 

15 Sret The in-lake retention of sulfur Estimated 

16 Nupt The net long-term average uptake flux of N in 

the biomass 

USFS-FIA data 

17 Nimm Immobilization of N in the soils Estimated fix value 

18 Nden Denitrification Estimated fix value 

19 Lake Size Lake size (ha) DLMs 

20 WSH Watershed area (ha) Calculated 

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The MAGIC model (Cosby et al., 1985a; 1985b; 1985c) is a mathematical model 

(a lumped-parameter model) of soil and surface water acidification in response to 

atmospheric deposition based on process-level information about acidification. A process 

model, such as MAGIC, characterizes acidification into (1)a section in which the 

concentrations of major ions are assumed to be governed by simultaneous reactions 

involving SO4

2- adsorption, cation exchange, dissolution-precipitation- speciation of 

aluminum, and dissolution-speciation of inorganic carbon; and (2) a mass balance section 

in which the flux of major ions to and from the soil is assumed to be controlled by 

atmospheric inputs, chemical weathering, net uptake and loss in biomass and losses to 

runoff. At the heart of MAGIC is the size of the pool of exchangeable base cations in the 

soil. As the fluxes to and from this pool change over time owing to changes in 

atmospheric deposition, the chemical equilibria between soil and soil solution shift to 

give changes in surface water chemistry. The degree and rate of change of surface water 

acidity thus depend both on flux factors and the inherent characteristics of the affected 

soils. 

 There are numerous input data required to run MAGIC making it rather data 

intensive. Atmospheric deposition fluxes for the base cations and strong acid anions are 

required as inputs to the model. These inputs are generally assumed to be uniform over 

the catchment. The volume discharge for the catchment must also be provided to the 

model. In general, the model is implemented using average hydrologic conditions and 

meteorological conditions in annual simulations, i.e., mean annual deposition, 

precipitation and lake discharge are used to drive the model. Values for soil and surface 

water temperature, partial pressure of carbon dioxide and organic acid concentrations 

must also be provided at the appropriate temporal resolution. 

The aggregated nature of the model requires that it be calibrated to observed data 

from a system before it can be used to examine potential system response. Calibrations 

are based on volume weighted mean annual or seasonal fluxes for a given period of 

observation. The length of the period of observation used for calibration is not arbitrary. 

Model output will be more reliable if the annual flux estimates used in calibration are 

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based on a number of years rather than just one year. There is a lot of year-to-year 

variability in atmospheric deposition and catchment runoff. Averaging over a number of 

years reduces the likelihood that an “outlier” year (very dry, etc.) is used to specify the 

primary data on which model forecasts are based. On the other hand, averaging over too 

long a period may remove important trends in the data that need to be simulated by the 

model. 

The calibration procedure requires that stream water quality, soil chemical and 

physical characteristics, and atmospheric deposition data be available for each catchment. 

The water quality data needed for calibration are the concentrations of the individual base 

cations (Ca, Mg, Na, and K) and acid anions (Cl, SO4

2-, and NO3-

) and the pH. The soil 

data used in the model include soil depth and bulk density, soil pH, soil cation-exchange 

capacity, and exchangeable bases in the soil (Ca, Mg, Na, and K). The atmospheric 

deposition inputs to the model must be estimates of total deposition, not just wet 

deposition. In some instances, direct measurements of either atmospheric deposition or 

soil properties may not be available for a given site with stream water data. In these 

cases, the required data can often be estimated by: (a) assigning soil properties based on 

some landscape classification of the catchment; and (b) assigning deposition using model 

extrapolations from some national or regional atmospheric deposition monitoring 

network. Soil data for model calibration are usually derived as aerially averaged values 

of soil parameters within a catchment. If soils data for a given location are vertically 

stratified, the soils data for the individual soil horizons at that sampling site can be 

aggregated based on horizon, depth, and bulk density to obtain single vertically 

aggregated values for the site, or the stratified data can be used directly in the model. 

A.1.3 Example of the two ways to calculate NECO from the first draft NOx and SOx 

secondary NAAQS Policy Assessment Document 

The steady-state critical load model suggest for use in the NAAQS by the PA 

could be constrained by a quantity of N which would be taken up, immobilized or 

denitrified by ecosystems and used to adjust the quantity of deposition required to meet a 

specified critical load. This term is abbreviated by Neco, and could be derived multiple 

September, 2010 Draft – Do Not Quote or Cite A-11
ways. The first is by taking the mean value calculated to represent the long-term amount 

of N an ecosystem can immobilize and denitrify before leaching (i.e. N saturation) that is 

derived from the FAB model. This approach requires the input of multiple ecosystem 

parameters. Its components are expressed by eq 1. 

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   Neco ret 1 imm  NNrNfNupt den 6 (1) 

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Where, 

Nupt= nitirogen uptake by the catchment 

Nimm= nitrogen immobilization by the catchment soil 

Nden=denitrification of nitrogen in the catchment, 

Nret = in-lake retention of nitrogen 

f =forest cover in the catchment (dimensionless parameter) 

r = fraction lake/catchment ratio (dimensionless parameter) 

The second approach for estimating Neco is to take the difference between N 

deposition and measured N leaching in a catchment as expressed by eq 2. 

eco  NNDLN leach ( ) (2) 

The site specific values of critical loads can be used to derive such a deposition 

loading, here called the deposition metric, which represents a group or percentage of 

water bodies that reach a specified ANC (or higher) in a given spatial area. For example, 

if it is desired that all water bodies reach a specified ANC, the allowable amount of 

deposition for all water bodies is equal to the lowest critical load the population of water 

bodies. Because the deposition metric represents a percentage of individual catchments 

from a population of water bodies, and not an individual catchment, the deposition metric 

is noted by the follow abbreviation DL%ECO. 

As an example of the above approach, we evaluated the population of 169 

waterbodies in the Adirondacks used in the REA analysis. For each individual waterbody 

in the population the critical load at ANClim = 50 μeq/L was calculated using the two 

equations for deriving the NECO term (eq 2 and 3). The distribution of deposition loads 

September, 2010 Draft – Do Not Quote or Cite A-12
for the population was assessed and Table A.1.3-1 shows example for the percentage of 

protection. The mean value for DL%ECO for the 169 water bodies is presented, as well as 

the values for which 50, 75, 85, 95 and 100% of the water bodies in the population will 

not exceed their critical load at ANC = 50 μeq/L. Note, only 32% of water bodies would 

not exceed their critical load at ANC = 50 μeq/L for the mean value DL%ECO because 

variability is high in the data set, therefore the mean can be problematic for areas with 

high variability. 

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Table A.1.3-1. Example calculations for determining the percent of water bodies achieving 

target ANC levels. This example is based the population of DLANClim for and ANC=50 for 169 catchments in 

the Adirondacks. These catchments occur across three categories of geologic sensitivity. We could separate 

the DLANClim values into sensitivity categories (if info is available) and do the analysis for each category or 

calculate one DLANClim for combined geologic categories. Units are in meq/m2

/yr. 

NHx 

dep 

Neco 

(eq2) 

DL%ECO (S+N) +Neco 

(eq 2) 

Neco 

(eq3) 

DL%ECO (S+N) 

+Neco (eq 3) 

% of lakes within the 

population that have 

ANC ≥ 50μeq/L 

Mean 20.40 19.19 162.36 63.95 207.55 31.7% 

St dev 3.22 3.03 162.92 11.15 165.42 

St er 0.25 0.23 13.04 0.86 13.24 

Rank 

%tile 

50% 99.33 139.22 50% 

75% 65.62 110.37 75% 

85% 54.89 95.53 85% 

95% 45.12 83.99 95% 

100% 30.22 59.07 100% 

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The values for the maximum deposition N and S based on DL%ECO at ANC=50 

using the two approaches for NECO and protective of 95 and 50% of the population of 

water bodies, are given in Table A.1.3-2. 

September, 2010 Draft – Do Not Quote or Cite A-13
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Table A-3 Values for N and S deposition tradeoff curves for ANC = 50, protecting 

95 and 50% of the population, in Adirondacks case study area as illustrated on Fig 

5.1 and Fig 5.2. Units are in meq/m2/yr unless noted otherwise. 

% protection NHxdep Neco DL%ECO (max N) DL%ECO (max S) DL%ECO (max NOY) 

95 Eq 2 20.4 19.19 45.12 25.93 141.96 

50 Eq 2 20.4 19.19 99.33 80.14 78.9.3 

95 Eq 3 20.4 63.75 83.99 20.04 63.59 

50 Eq 3 20.4 63.75 139.22 75.27 118.82 

A.1.4 Additional Analysis of Bedrock Geology as a basis for defining acid-sensitivity 

classes across the landscape 

APPROACH 

This analysis applied a methodology developed by Sullivan et al. (2007) for their 

exercise to test the hypotheses on watershed sensitivity to acidic deposition. Sullivan et 

al.’s exercise focused on streams in the Southern Appalachian Mountains region. The 

classification scheme was based on lithologic maps and the stream chemistry for more 

than 900 sites. Using logistic regression to model the presence of acid-sensitive 

waterbodies (expressed as ANC) with respect to lithology class, the authors found “four 

variables were highly significant in predicting the probability of occurrence of acidsensitive sites (defined for this analysis as having ANC < 20 μeq/L) in the Southern 

Appalachians using logistic regressions: % siliceous bedrock in watershed, % forested 

watershed, elevation, and watershed area”. (Sullivan et al., 2007) 

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Using stream and lake ANC data from the Shenandoah Mountains and the 

Adirondack Mountains, along with lithologic classifications, linked forested area, 

elevation and watershed area, this NOx SOx exercise used logistic regression to 

determine if a correlation exists between lithology and ANC that is similar to the 

correlation found by Sullivan et al. in the Southern Appalachians. If similar to Sullivan 

et al.’s findings, then the findings of this NOx SOx exercise will be extrapolated spatially 

to identify areas of the U.S. potentially sensitive to aquatic acidification. 

September, 2010 Draft – Do Not Quote or Cite A-14
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Phase I of this analysis applied the Sullivan et al. methodology separately to the two 2009 NOxSOx Risk and Exposure Assessment (Shenandoah Mountains and the Adirondack Mountains 

Case Study Areas) and consists of 4 steps: 

6 1. Site selection (i.e., lakes and streams) 

7 2. Stream chemistry data acquisition and conversion to a spatial data layer 

8 3. Data acquisition for % lithology classification of watershed (e.g., % siliceous bedrock), 

elevation, % forested area of watershed, and watershed area 

4. Logistic regression modeling. 

The modeling results are compared to Sullivan’s SAMI analysis and, if consistent, 

will be extrapolated to the region surrounding the case study area (Phase II) and to other 

regions of the U.S. (Phase III) upon EPA directive. 

The output desired for this Phase I exercise was a set of constants and coefficients 

that predict the probability that the acid-sensitive watersheds (having stream ANC either 

< 0 μeq/L or between 0 and 20 μeq/L) in the selected study areas based upon the input 

variables (percent lithology classification in watershed, elevation, watershed size, and 

percent forest cover in watershed). 

PHASE I 

Step 1 – Site Selection: Streams and lakes to be assessed 

The Adirondack Case Study Area and the Shenandoah Case Study Area provide ideal 

areas to assess the risk to aquatic ecosystems from NOx and SOx acidifying deposition. Four 

main reasons support the selection of these two areas. First, both regions fall within the areas of 

the United States known to be sensitive to acidifying deposition because of a host of 

environmental factors that make these regions predisposed to acidification. Second, these areas 

are representative of other areas sensitive to acidification. Third, these regions have in the past 

and continue to experience substantial exposure to NOx and SOx air pollution. Fourth, these 

September, 2010 Draft – Do Not Quote or Cite A-15
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areas have been extensively studied ... over the last 3 decades (see Section 4 of the ISA Report 

(US EPA 2008). (US EPA, 2009, REA, Appendix 4, Section 3.1) 

Freshwater surveys and monitoring in the eastern United States have been 

conducted by many programs since the mid-1980s, including the National Lake/Stream 

Surveys (NSWS), EPA’s Environmental Monitoring and Assessment Program (EMAP), 

the Temporally Integrated Monitoring of Ecosystems (TIME) monitoring program 

(Stoddard, 1990), and Long-Term Monitoring (LTM) project (Ford et al., 1993; Stoddard 

et al, 1998). The purpose of these programs is to determine the current state and 

document the trends over time in surface water chemistry for regional populations of 

lakes or streams impacted by acidifying deposition. Based on extensive surveys and 

surface water data from these programs, it was determined that the most sensitive lakes 

and streams (i.e., ANC less than about 50 μeq/l) in the eastern US are found in New 

England, the Adirondack Mountains, the Appalachian Mountains (northern Appalachian 

Plateau and Ridge/Blue Ridge region), northern Florida, and the Upper Midwest. These 

areas are estimated to contain 95% of the lakes and 84% of the streams in the United 

States that have been anthropogenically acidified through deposition (see Annex 4.3.3.2 

of the ISA, US EPA 2008). (US EPA, 2009, REA, Appendix 4, Section 3.1) 

ANC in surface water from 50 lakes in the Adirondack Case Study Area were monitored 

through the Adirondack LTM program and 38 lakes from the TIME program (Figure A.1.4-3).

For the Adirondack Case Study Area, the regional EMAP probability survey of 117 lakes 

(i.e., weighting factors) were used to infer the number of lakes and percentage of lakes that 

receive acidifying deposition above their critical load of a target population of 1,842 lakes. The 

117 lakes...represent a subset of 344 sampled lakes throughout New England from 1991 through 

1994. (ME, NH, VT, RI, MA, CT, NY, NJ). (U.S. EPA, 2009, Appendix 4, Section 4.3.1.1) 

Shenandoah Case Study Area ANC monitoring occurred as part of the 

Shenandoah National Park Surface Water Acidification Study (SWAS), Virginia Trout 

Stream Sensitivity Survey (VTSSS), and LTM programs. There are a significant number 

of the 67 streams in SWAS-VTSSS and LTM programs that currently have ANC of 

September, 2010 Draft – Do Not Quote or Cite A-16
about 50 μeq/L based on the observed annual average ANC concentrations (Figure 

A.1.4-4). 

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The total number of brook trout streams represented by the SWAS-VTSSS LTM 

quarterly monitored sites is approximately 310 out of 440 mountain headwater streams known to 

support reproducing brook trout in the Shenandoah Case study Area. ... The SWAS-VTSSS LTM 

programs began in Spring 1987, when water samples were collected form 440 streams known to 

have brook trout. Following the 1987 survey , a representative subset of 69 streams was selected 

for long-term quarterly monitoring of water quality, mostly located on National Forest lands or 

within the Shenandoah National Park Case Study Areas (14 SWAS and 55 VTSSS streams). 

(U.S. EPA, Appendix 4, Section 4.3.1.2). 

Figure A.1.4-3. Current yearly average for 2005 to 2006 ANC (μeq/L) in 

surface waters from 88 monitored lakes in the Temporally Integrated 

Monitoring of Ecosystems (38 Lakes) and Adirondacks Long-Term 

Monitoring (50 Lakes) networks in the Adirondack Case Study Area. 

(U.S. EPA, 2009) 

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September, 2010 Draft – Do Not Quote or Cite A-17
Figure A.1.4-4. Current yearly average for 2005 to 2006 ANC (μeq/L) in surface 

waters from 67 monitored streams in the Surface Water Acidification Study, 

Virginia Trout Stream Sensitivity, and Long-Term Monitoring network in the 

Shenandoah Case Study Area. 

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[In Phase I, efforts focused on the Shenandoah and Adirondack Mountains case study 

areas.] 

Step 2: Stream Chemistry Data Acquisition - Water chemistry for those streams and lakes

 

There are numerous chemical constituents in surface water that can be use to indicate the 

acidification condition of lakes and streams and to assess the effects of acidifying deposition on 

ecosystem components. These include surface water pH and concentrations of SO4

2-, NO3-

, Al3

+

, 

and Ca2

+

; the sum of base cations; the recently developed base cation surplus; and the acid 

neutralizing capacity (ANC). Each of these chemical indicators provides direct links to the health 

of individual biota and the overall health and integrity of aquatic ecosystems as a result of surface 

water acidification. 

Although ANC does not directly affect the health of biotic communities, it is calculated 

(or measured) based on the concentrations of chemical constituents that directly contribute to or 

ameliorate acidity-related stress, in particular, pH, Ca2

+

, and dissolved inorganic aluminum. 

Furthermore, numerical models of surface water acidification can more accurately estimate ANC 

than all of the individual constituents that comprise it. Consequently, for the purpose of the 

performing the case studies reported in the Risk and Exposure Assessment (US EPA, 2009), 

September, 2010 Draft – Do Not Quote or Cite A-18
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annual average ANC of surface waters was used as the primary metric to quantify the current 

acidic conditions and biological impacts for a subset of waterbodies in the study areas. (U.S. 

EPA, 2009 (REA, Appendix 4, Section 2.2)). 

All lake and stream monitoring data within the Adirondack and Shenandoah study 

areas were used.. Although Sullivan et al. used only Spring data because those values 

represented the time of year when ANC is at its lowest, and is also a time when sensitive 

life stages (eggs and young) are present for many fish species (Kaufman, Herlihy, Mitch, 

Messer,& Overton, 1991) this study used data throughout the calendar year. This was 

done because of the considerable variation within and between the study areas in terms of 

the time of lowest seasonal ANC, and the life cycles of the dependent biota. ANC 

values were then averaged by sampling station. Those stations in the Adirondacks and 

Shenandoah regions that fell within a 12 digit HUC that crossed into either of the study 

area boundarieswere extracted to separate GIS data layers. These two collections of 12 

digit HUCs were then used to extract and average the other data parameters. 

Step 3: Data acquisition for percent lithology classification, elevation, percent forested area 

of watershed, and watershed size 

Bedrock and surficial lithology – Information on lithology for the eastern United States 

was been acquired from the states’ Geologic Surveys (USGS) and processed by EPA. This 

information was needed to determine if Sullivan et al.’s geologic classification system for the 

Southern Appalachian Mountains as applied to the eastern United States is a significant predictor 

of acid sensitivity in the Shenandoah Mountains and the Adirondack Mountains. Determining 

which acid class a given rock lithology falls into can be a somewhat subjective process. Upon 

initial inspection of the data, it appeared that there were some differences found between adjacent 

states (Figure A.-1). Following discussion with EPA, it was determined that the Adirondack 

data would benefit from re-evaluation of the classifications assigned. The acid sensitivity classes 

were reviewed for consistency and correctness and found to be accurate in all but a handful of 

cases. These exceptions were modified to match their lithologic descriptions. The Shenandoah 

class assignments were deemed acceptable. The percent of each of the five classes of acid 

sensitivity (siliceous, argillaceous, felsic, mafic, and carbonate) were tabulated over each of the 

12 digit HUCs that fell within the study areas.

September, 2010 Draft – Do Not Quote or Cite A-19
Figure A.1 Distribution of geology acid classes in the eastern United States and the 

location of the ELSDS4 sample locations. 

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Elevation, feet (30m Digital Elevation Maps) The water quality sample locations for the 

Adirondack and Shenandoah Mountains were extracted to a single shapefile, which was then 

overlaid with a national 30 meter digital elevation model (DEM). The elevation was determined 

at each location and extracted as an attribute. 

Watershed area, km2

 (HUC12) The HUC 12 identifier was extracted using GIS in a 

manner similar to the extraction of elevation information for each acid-sensitive sample location. 

The area (km2

) of the HUC 12 watershed was extracted to the water sampling location shapefile 

as an attribute. 

Percent forested area by watershed has been estimated using 1992 US Forest Service 

vegetative databases and the Advanced Very High Resolution Radiometer (AVHRR) satellite 

September, 2010 Draft – Do Not Quote or Cite A-20
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data (1,000 m cell size) to produce a raster forest type image with a 250 m cell size. This forest 

type image was processed using ArcGIS’s Spatial Analyst by HUC 12 to determine the number 

of forested versus non-forested cells in each watershed. The results of this analysis were then 

used to calculate the percent of forested area in each watershed. 

Step 4: Logistic Regression Modeling 

Following Sullivan et al. (2007) a logistic regression approach with stepwise 

variable selection procedure was used to model binary variables derived from ANC. A 

significance level of = 0.20 was used as the criterion for variable entry, and = 0.25 

was selected as the criterion for remaining in the model (Hosmer and Lemeshow (1989), 

p.108). The following continuous predictor variables were considered in this modeling 

exercise : Elevation (ft), Area (km2

), % Felsic, %_Carbonate, %Mafic, %_Siliceous, % 

Argillaceous, and %_FOREST. SAS PROC Logistic (SAS (version 9.2)) was used to 

determine the best model. 

Three binary responses were considered in the logistic modeling: 

1-. ANC<=20 then outcome=1; else outcome=0 

2- ANC<=50 then outcome=1; else outcome=0 

3- ANC<=100 then outcome=1; else outcome=0. 

ADDITIONAL INFORMATION ABOUT THE METHODOLOGY 

The model was run for all sites’ elevations and then followed by subsequent 

model runs that eliminate sites below a selected elevation and above a selected elevation 

to discern any improvement in correlation sensitive to elevation. The lower and upper 

elevations were selected based on the distribution of locations provided in Step 1. 

EXTRAPOLATION – PHASES II AND III 

The overall goal of this exercise was to develop constants and coefficients that 

predict the acid-sensitivity of watersheds with a high degree of probability. A different 

set of coefficients were developed for each distinct geographic region for which there is 

September, 2010 Draft – Do Not Quote or Cite A-21
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water quality monitoring data. In Phase II, the acid-sensitivity probability was calculated 

and mapped using GIS for all watersheds within each distinct geographic region of the 

case study areas, using the coefficients developed within the region. 

In Phase III, the acid-sensitive probability for watersheds in other regions outside 

the sampled regions would be calculated and mapped using GIS by applying the set of 

coefficients from the sampled region which is most similar, and for which all 4 types of 

input data are available (i.e., percent lithology classification of watershed, percent 

forested watershed, elevation, and watershed area). Classified lithology data are 

currently only compiled for the eastern half of the United States, so extrapolation will 

only be possible for this portion of the country at this time. 

RESULTS 

GIS Mapping of Geologic Classifications and Measured ANC 

Figures A-2andA-3 show the distribution of water sample locations in the Adirondack 

and Shenandoah Mountains, respectively. The average ANC values are symbolized 

using 20, 50, and 100 microequilavlents/litre as the break points. The five geology acid 

classes are also shown. 

September, 2010 Draft – Do Not Quote or Cite A-22
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2 Figure A-2. Average ANC and Geology Classification: Adirondack Region 

September, 2010 Draft – Do Not Quote or Cite A-23
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Figure A-3. Average ANC and Geology Classification: Shenandoah Region 

Logistic Regression Analysis Results 

Before running any modeling technique, the correlation between the predictors and ANC 

was explored (See Figure A.-4) A cell in the interception of a row and column represents 

the correlation between the variables at the end of the row and column. The correlation 

coefficient, R2

 , ranges from -70 to 30%, suggesting a low to moderate correlation 

between predictors and outcome. Negative correlations indicate that as the predictor 

value increases, the value of ANC decreases. Positive correlations suggest that as the 

value of the predictor increases so does the value of ANC. The correlation matrix also 

serves to explore the multi-co-linearity issue (presence of highly correlated predictors). 

In the Adirondacks region, Elevation is the variable with higher correlation (-30% to -

50%), suggesting that this variable will likely be included as a predictor in the model. 

In the Shenandoah region, % Carbonate (R2

 between 40% and 70%) and % Forest (R2

between -40% to -70%) show a modest to high correlation with ANC. (See Figure A-5.) 

Also, % Forest is moderately correlated with % Siliceous (R2

 between 40% to 70%) and 

September, 2010 Draft – Do Not Quote or Cite A-24
% Carbonate (R2

 between -40% to -70%), suggesting that it is likely that one of these 

three variables will be in the model. Similarly, % Felsic is moderately correlated with % 

Carbonate and % Argillaceous (R2

 between -40% to -70%) suggesting that if one of these 

variables is in the model, then the other two won’t be in the model. 

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September, 2010 Draft – Do Not Quote or Cite A-25
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Among the measurements of model assessment considered were C statistic (a C value of 

0.5 means the model predictions are equivalent to a random guess; >0.7 denotes 

worthwhile to the model); the Akaike information Criteria (AIC) (Hosmer and 

Lemeshow (1989), p.184); and the Schwarz information criterion (SC) (Kass, R. E. & 

Wasserman, L. (1995)); and the -2*maximized log-likelihood (-2ML). Smaller values of 

all three statistics suggest a better model. To evaluate the goodness of fit of the model, 

the Hosmer and Lemeshow goodness of fit test (Hosmer and Lemeshow (1989)) was 

used (larger p-values suggest no difference between observed and model-predicted 

values, implying that the model's estimates fit the data at an acceptable level). 

September, 2010 Draft – Do Not Quote or Cite A-26
1 The final model for the Adirondacks Region binary variables is the following: 

2 

0

0

For the binary variable 20

1 2

0 2

ANC

Y

ANC

   

 

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the logistic model is the following: 

logit( ) 3.3933 0.00305 Elevation 0.0142 (%Forest) 0.0216 (%Carbonate)

 0.0366 (%Siliceous) 0.0172 (%Mafic)

 

p       

 

For the binary variable 50

1 5

0 5

ANC

Y

ANC

   

 

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logit( ) 4.2271 0.00306Elevation 0.0197(%Carbonate) 0.0463(%Siliceous) 0.0155(%Mafic)

 

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For the binary variable 100

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0 100

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ANC

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 

the logistic model is the following: 

The final model for the Shenandoah region binary variables is the following: 

For the binary variable 20

1 2

0 2

ANC

Y

ANC

   

 

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the logistic model is the following: 

logit( ) 5.8725 0.0358(%Carbonate)+0.0632(%Siliceous)

 

p  

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ANC

Y

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the logistic model is the following: 

logit( ) 6.0745 0.0233(%Carbonate)+0.0750(%Siliceous)

 

p  

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0 100

ANC

Y

ANC

   

 

September, 2010 Draft – Do Not Quote or Cite A-27
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the logistic model is the following: 

logit( ) 3.9068 0.000699(%Carbonate)+0.047(%Siliceous)

 

p  

Betas, standard error for betas, odds ratio and relevant statistics can be found in Appendix 

2. 

Weibull modeling using Proc Reliability (SAS) was also used to model ANC. LCL and 

UCL denote the lower and upper confidence limits at 95% confidence level. If the value 0 

is contained between the confidence limits then the parameter is not statistically 

significant at 5% level. 

For the Adirondacks Region, the final predictors and estimates are shown below: 

 Parameter Estimate SE 95% LCL 

95%UCL 

            Intercept               7.9493        0.3694        7.2254        8.6733

            ELEV_FT                ‐0.0015        0.0001       ‐0.0016       ‐0.0013

            PCT_FOREST             ‐0.0085        0.0039       ‐0.0162       ‐0.0008

            PCT_Carbonate           0.0021        0.0012       ‐0.0003        0.0045

            PCT_Argillaceous        0.0533        0.0208        0.0125        0.0941

According to this model, ANC decreases with each unit increase in Elevation and Forest 

(negative estimates) and increases with each unit increase in % Carbonate and % 

Argillaceous (positive estimates). 

Model assessment statistics (R2, AIC, etc.) were not found in the documentation. It is 

possible to calculate them to evaluate the model fit to the data. 

For the Shenandoah Region, the final predictors and estimates are shown below: 

            Intercept               7.3437        0.1585        7.0330        7.6543

            ELEV_FT                ‐0.0015        0.0001       ‐0.0016       ‐0.0013

            PCT_Siliceous           0.0167        0.0054        0.0061        0.0273

            PCT_Argillaceous        0.0564        0.0209        0.0154        0.0973

            PCT_Felsic             ‐0.0041        0.0011       ‐0.0062       ‐0.0020

According to this model, ANC decreases with each unit increase in Elevation and % 

Felsic (negative estimates) and increases with each unit increase in % Siliceous and % 

Argillaceous (positive estimates). 

Appendix 2 shows the odds ratio for each predictor. The larger the Odds ratio, the 

higher the chances that ANC <=20 for every change in unit of the predictor. 

September, 2010 Draft – Do Not Quote or Cite A-28
1 For the Adirondacks region and ANC<=20 

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33 

When elevation increases one unit, the odds that ANC<=20 increase by a factor of 1.03, 

when other variables are controlled (meaning the same level of % Fforest, % Carbonate, 

% Siliceous and % Mafic). When % Forest increases one unit, then the odds that 

ANC<=20 decreases by a factor of 0.972 when other predictors are held constant. Similar 

interpretations exist for the rest of the models. 

For the Shenandoah region and ANC<=20: 

When % Carbonate increases one unit, the odds that ANC<=20 increases by a 

factor of 1.036 when other predictors are held constant. Similarly,when % Siliceous 

increases one unit, the odds that ANC<=20 increases by a factor of 1.065 when other 

variables are held constant. 

UNCERTAINTIES 

Lithology classification assignments. RTI received a geodatabase of geology data 

from EPA to be used in the ANC GIS comparative analysis. The geodatabase includes a 

list of lithologies from the most recent USGS (Eastern Mineral Resources Team) national 

geologic map data for the continental United States. Each lithology polygon from the 

USGS map data set is defined by two lithology types labeled’ROCKTYPE1’ and 

‘ROCKTYPE2’. According to the metadata for the geologic map cover, ROCKTYPE1 

characterizes more than 50% of the area in the associated polygon. EPA has categorized 

each lithology record into one of five geologic sensitivity classes (i.e., Siliceous, 

Argillaceous, Felsic, Mafic, and Carbonate), after Sullivan, et al (2006). The geologic 

sensitivity classification is included for each lithology record of the geodatabase in a field 

indentified as ‘AcidClassDesc’. 

A cursory review was performed on EPA’s assigned classifications suggesting-- 

 Not all rock types in the geodatabase are included in Sullivan (Table 1), but in 

most instances (except where noted) there was agreement between Sullivan’s 

classification and the EPA assigned classes for rock types included in both the 

technical document and the geodatabase; 

September, 2010 Draft – Do Not Quote or Cite A-29
1  Potential discrepancies in class assignments were identified such as those 

associated with non-descriptive lithogic description (e.g., sedimentary rock), 

others with typical rock chemistry and GSC (e.g., Carbonate GSC was assigned to 

beach sand), also classification assigned to sedimentary environments (e.g., 

olistostrome), or terms describing non-mineralogy specific rock texture (e.g., 

hornfel). The majority of these fell outside the Adirondack or Shenandoah study 

areas. 

2 

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4 

5 

6 

7 

8 

9 

10 

11 

12 

After discussion with EPA and a more detailed review, only a few revisions were 

required in the Adirondack datasets and those revisions were assigned less acid-sensitive 

classes. No revisions were required in the Shenandoah dataset. 

 

September, 2010 Draft – Do Not Quote or Cite A-30
1 

2 

3 

4 

5 

6 

7 

A.1.5 Additional techniques explored for which to base subdivision of the U.S. based based 

on acid sensitivity. 

Alkalinity 

We also considered dividing the U.S. based on the Alkalinity from the 1984 

Omernick Alkalinity map. We do not recommend using alkalinity because it includes 

more ecoregions as sensitive than the ANC approach, which causes the deposition loads 

for sensitive and less-sensitive areas to be more similar.

Fig A.1.5-1. Omernick’s alkalinity map (Omernick 1984) 

8 

September, 2010 Draft – Do Not Quote or Cite A-31
1 

Table A.1.5-1. Descriptive statistics of the CL populations that result when the U.S. is divided 

into two categories, sensitive and non-sensitive based on Alk 

Omernick sensitivity classification

CL Example deposition metrics 

for 

ANC

= 

Sensiti

vity 

class 

CL in class 

n=

Mean 

Neco 

Neco St 

er 

CL mean 

(meq/m2/yr) 

CL 

St er 

DL %90 

(meq/m2/yr) 

DL %75 

(meq/m2/yr) 

DL %50 

(meq/m2/yr) 

50 sensit

ive 

5446 48 0.4 274 5 28 57 123 

50 Nonsensit

ive 

346 37 1 343 15 50 109 265 

2 

3 

4 

5 

6 

7 

8 

Cluster Analysis 

This approach uses ANC, ALK, DOC and soil BCw to categorize the acidsensitivity of ecoregions based on a quantitative cluster analysis. The ANC dataset is 

described in Chapter 5, the ALK dataset is described in the previous section, and the 

DOC and soil BC weathering datasets are described prior to the explanation of the cluster 

analysis

9 DOC

10 

11 

12 

13 

Water chemistry data on DOC was collected from several national monitoring 

networks. The data sources are the same as those for ANC and summarized in Chapter 

5and include approximately 11,000 observations. 

14 Bed rock geology

15 

16 

17 

18 

19 

20 

21 

22 

23 

Bedrock geology is the parent material to soils and can be related to acidsensitivity. In the first draft PA the EPA staff proposed using bedrock geology to inform 

acid sensitivity categorization of the landscape. This approach entailed defining 

categories of ecosystem sensitivity to acidification based on bedrock geology/ lithology. 

The approach is supported by conclusions from the ISA in which geology is determined 

to be the governing factor that drives ecosystem sensitivity to acidification (ISA 3.2.4.1). 

A map was developed in this policy assessment to capture the heterogeneity of geologic 

bedrock that occurs across the eastern U.S. and link it to ecosystem acid-sensitivity (Fig 

A.1.5-2). The method is based on Sullivan et al. (2007) in which 70+ primary lithologies 

September, 2010 Draft – Do Not Quote or Cite A-32
1 

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3 

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10 

11 

12 

13 

are grouped into 5 categories of acid-sensitivity, using ANC as the ecosystem indicator 

upon which acid-sensitivity is based. Sullivan et al. (2007) evaluated multiple features of 

the landscape and found that geology is the landscape parameter that governs ecosystem 

sensitivity to acidic deposition. The analysis in Sullivan et al. 2007 was conducted in the 

Southern Appalachian Mountains region, which included sites from the states of GA, TN, 

NC, KT, VA and WV. EPA conducted additional analyses to further test the concept that 

lithology correlates to acid sensitivity based on ANC, using data from the Adirondacks 

and Shenandoahs. Results from this analysis indicated that North of the glaciation line, 

bedrock geology was not a good predictor of acid sensitivity. For example, the 

Adirondacks which are known to be one of the most sensitive areas to acidification is 

dominated by types of bedrock geology that are considered moderately to less sensitive in 

areas South of the glaciation line. 

September, 2010 Draft – Do Not Quote or Cite A-33
1 

2 

3 

4 

Fig A.1.5-2. A map of acid sensitive areas of the Eastern U.S. 

developed from a lithology-based five-unit geologic classification 

system after methods in Sullivan et al. (2007). 

5 Soil base-cation weathering

6 

7 

8 

9 

As previously noted, the use of bedrock geology for the classification of acid sensitive 

catchments is problematic in areas north of the glaciation line because glacial activity has 

caused the surface till to become spatially disconnected from the parent geologic material. 

Surficial geology is therefore important to the evaluation of the spatial variability of acidSeptember, 2010 Draft – Do Not Quote or Cite A-34
1 

2 

3 

sensitivity. For that reason a national map for soil base-cation weathering (McNulty et al. 

2007) was considered in the analysis of acid-sensitivity (Fig A.1.5-3). 

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11 

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15 

Fig A.1.5-3. A map of average annual forest soil base cation 

weathering expressed in eq ha-1 yr-1 for the conterminous US for the 

years 1994 to 2000 at a 1-km2 spatial resolution from McNulty et al. 

McNulty et al. 2007 estimated the base cation weathering rate using the clay correlationsubstrate method (Sverdrup et al., 1990). This method used a combination of parent 

material and clay percent to determine the weathering rate. Base cation weathering rate 

(eq ha-1

 yr-1). A temperature correction can be applied to this method, but this correction 

is more suitable for northern climates and was not used in the model. Clay fraction was 

derived from a weighted average of soil fraction per map unit. 

September, 2010 Draft – Do Not Quote or Cite A-35
1 

2 

Table A.1.5-3. Data input to the U.S. national soil base cation weathering map 

developed by McNulty et al. 2007

Data source Description of dataset 

State Soil 

Geographic Data 

Base (STATSGO)

Compilation by the Natural Resource Conservation Service (NRCS) of geology, 

topography, vegetation, climate, Landsat Thematic Mapper (TM) satellite imagery, 

and detailed, county level soil survey data. Each soil map unit in STATSGO includes 

multiple components and data layers (USDA NRCS, 1995). STATSGO is organized 

by state and consists of one geospatial vector representing the soil map units for that 

state and 15 tables describing characteristics of those map units. Multiple soil layers 

are associated with each map unit. Soil layer sampling depth is not consistent within a 

state or between states

CONUS-SOIL 

developed by the 

Earth System 

Science Center 

(ESSC) at 

Pennsylvania 

State University

This is a 1-km2

 multi-layer soil data set based on the STATSGO. ESSC converted the 

vector map unit layer in STATSGO to a 1-km2

 grid, remapped many of the original 

STATSGO attribute layers, and defined 11 standard soil layers (Miller and White, 

1998). These data layers and tables linked the standardized data set to the original 

STATSGO data set distributed by ESSC as 1-km2

 soil map unit grids for the 

conterminous US. The CONUS-SOIL was much better suited for national-scale 

modeling than the original STATSGO attribute layers and was therefore used 

McNulty et al. 2007. Key soil data inputs included CONUS-SOIL map units and clay 

fraction (Miller and White, 1998).

forest soil percent 

organic matter 

(OM) layer

Was created by McNulty et al. 2007 by first averaging the maximum and minimum 

recorded values for OM were averaged for each layer in the STATSGO data set. 

Second, the average OM layers were remapped into the 11 standard CONUS-SOIL 

layers using a weighted average to redistribute the average OM STATSGO layers into 

the CONUS-SOIL layers. If a STATSGO layer was completely contained in a 

CONUS-SOIL layer, then the average OM was multiplied by the component percent 

to determine the average OM contribution to the standard layer. If the STATSGO 

layer overlapped more than one CONUS-SOIL layer, then the proportion of overlap 

was multiplied by the average OM and the component percent, where the component 

percent was the proportion of the soil map unit comprised of that soil component. 

Once the conversion from STATSGO to CONUS-SOIL layer was complete, the 11 

standard layers were summed by layer and divided by the sum of component percent. 

Finally, the weighted average was calculated according to equation 8. 

parent material 

class

was derived from the STATSGO map unit component (comp) and taxonomic (tax) 

classification tables (USDA NRCS, 1995). The dominant mineralogy for each soil 

map unit was determined from the comp and tax tables. Each unit was classified into 

parent material class based on the mineralogical description (USDA NRCS, 2006)

Soil depth in meters was obtained from the CONUS-SOIL depth to bedrock layer. This layer 

identified map units with bedrock less than 1.52 m below the soil surface (i.e., map 

units coded with a depth of 1.52 m did not encounter bedrock) (Miller and White, 

1998). 

3 

4 

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9 

Results 

This approach uses ANC, ALK, DOC and soil BCw to categorize the acid-sensitivity of 

ecoregions based on a quantitative cluster analysis. Cluster analysis is a method to sort a 

set of observations (in this case CL from watersheds) into subgroups so that the degree of 

association between observations is maximal if they belong to the same group and 

September, 2010 Draft – Do Not Quote or Cite A-36
1 

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minimal otherwise. The similarity of each site was determined using ANC, ALK, DOC 

and BCw. In the cluster analysis, the k-means algorithm assigned each point to the cluster 

whose center (also called centroid) is nearest. The center is the average of all the points in 

the cluster — that is, its coordinates are the arithmetic mean for each dimension 

separately over all the points in the cluster. The ecoregions was then assigned 

membership to a cluster. The weakness of this approach is that each of the datesets for 

the criteria (ANC, soilBCw, Alk and DOC), have varying levels of national coverage. If 

not all four criteria are available for the analysis the ecoregions drops out and is not 

assigned a cluster identification. The majority of ecoregions in the U.S. were not 

represented in the datasets for all four criteria and were not assigned a cluster 

identification (Fig A.1.5-4). Therefore the ulitility of this approach is limited. 

13 

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September, 2010 Draft – Do Not Quote or Cite A-37
1 REFERENCES 

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Kaufmann, P.E., A.T. Herlihy, M.E. Mitch, J.J. Messer, and W.S. Overton, 1991. 

Chemical characteristics of streams in the Eastern United States: I. Synoptic 

survey design, acid-base status and regional chemical patterns. Water Resources 

Research, 27, 611-627. 

NAPAP (National Acid Precipitation Assessment Program), 2005. National Acid 

Precipitation Assessment Program Report to Congress: An Integrated 

Assessment. National Acid Precipitation Assessment Program, Washington, DC. 

Stoddard, J., J.S. Kahl, F.A. Deviney, D.R. DeWalle, C.T. Driscoll, A.T. Herlihy, J.H. 

Kellogg, P.S. Murdoch, J.R. Webb, and K.E. Webster, 2003. Response of 

Surface water Chemistry to the Clean Air Act Amendments of 1990. EPA 620/R03-001. U.S. Environmental Protection Agency, Office of Research and 

Development, National Health and Environmental Effects Research Laboratory, 

Research Triangle Park, NC. 

Sullivan, T.J., C.T. Driscoll, B.J. Cosby, I.J. Fernandez, A.T. Herlihy, J. Zhai, R. 

Stemberger, K.U. Snyder, J.W. Sutherland, S.A. Nierzwicki-Bauer, C.W. Boylen, 

T.C. McDonnell, and N.A. Nowicki. 2006. Assessment of the Extent to which 

Intensively-Studied Lakes are Representative of the Adirondack Mountain 

Region. Final report. New York State Energy Research and Development 

Authority (NYSERDA), Albany, NY. Available at 

http://nysl.nysed.gov/uhtbin/cgisirsi/Qcwd6NzFby/NYSL/138650099/8/4298474 

(accessed November 1, 2007). 

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32 

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39 

40 

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43 

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Sullivan, T. J., J. R. Webb, K.U. Snyder, A.T. Herlihy, B.J. Cosby, 2007.“Spatial 

Distribution of Acid-sensitive and Acid-impacted Streams in Relation to 

Watershed Features in the Southern Appalachian Mountains”. Water Air Soil 

Pollution (2007) 182:57-71 

U.S. EPA, 2009. Risk and Exposure Assessment for Review of the Secondary National 

Ambient Air Quality Standards for Oxides of Nitrogen and Oxides of Sulfur. 

EPA-452/R-09-008b. Final. U.S. Environmental Protection Agency, Office of 

Air Quality Planning and Standards. Research triangle Park, NC. 

U.S. EPA, 2008. Integrated Science Assessment (ISA) for Oxides of Nitrogen and 

Sulfur-Ecological Criteria. Final Report. EPA/600/R-08/082F. U.S. 

Environmental Protection Agency, National Center for Environmental 

Assessment – RTP Division, Office of Research and Development, Research 

Triangle Park, NC. 

Davis, J. M., Swall, J. L., 2006. An examination of the CMAQ simulations of the wet 

deposition of ammonium from a Bayesian perspective. Atmospheric Environment 

40, 4562–4573. 

EPA, 1999. Science Algorithms of the EPA Models-3 Community Multiscale Air Quality 

(CMAQ) Modeling System. Tech. Rep. EPA/600/R-99/030, U.S. Environmental 

Protection Agency, Washington DC. 

Grantz, D., Garner, J., Johnson, D., 2003. Ecological effects of particulate matter. 

Environment International 29, 213-239. 

September, 2010 Draft – Do Not Quote or Cite A-38
September, 2010 Draft – Do Not Quote or Cite A-39

1 

2 

3 

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5 

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9 

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13 

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15 

16 

Lien L; Raddum GG; Fjellheim A. (1992). Critical loads for surface waters: invertebrates 

and fish. (Acid rain research report no 21). Oslo, Norway: Norwegian Institute for 

Water Research 

Morris, R. E., McNally, D. E., Tesche, T. W., Tonnesen, G., Boylan, J. W., Brewer, P., 

2005. Preliminary evaluation of the community multiscale air quality model for 2002 

over the Southeastern United States. Journal of the Air and Waste Management 

Association 55, 1694–1708. 

Seinfeld, J., Pandis, S., 1998. Atmospheric Chemistry and Physics. John Wiley and Sons, 

Inc., New York. 

Sullivan TJ; Webb JR; Snyder KU; Herlihy AT; Cosby BJ. (2007). Spatial distribution of 

acid-sensitive and acid-impacted streams in relation to watershed features in the 

southern Appalachian mountains. Water Air Soil Pollut, 182, 57-71. 

Sullivan TJ; Fernandez IJ; Herlihy AT; Driscoll CT; McDonnell TC; Nowicki NA; 

Snyder KU; Sutherland JW. (2006). Acid-base characteristics of soils in the 

Adirondack Mountains, New York. Soil Sci Soc Am J, 70, 141-152. 
1 APPENDIX B 

2

3

Analysis of Critical Loads, Comparing Aquatic and Terrestrial Acidification 

4 Background

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Critical load is defined as, “a quantitative estimate of ecosystem exposure to one or more 

pollutants below which significant harmful effects on specified sensitive elements of the 

environment do not occur, according to present knowledge” (McNulty et al., 2007), and critical 

loads can be estimated for aquatic and terrestrial ecosystems. Within the Risk and Exposure 

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

Nitrogen and Sulfur (hereafter referred to as REA Report) (US EPA, 2009), critical loads of 

acidification for aquatic systems were determined by relating specific amounts of acidifying 

nitrogen and sulfur deposition to selected Acid Neutralizing Capacities (ANC) within freshwater 

lakes or streams. The presence and abundance of fish species served as the biological indicator 

of the impacts of the exceedance of critical acid loads by nitrogen and sulfur deposition. 

Estimation of critical acid loads for terrestrial systems within the REA Report (US EPA, 2009) 

related acidifying nitrogen and sulfur deposition to the base cation to aluminum (Bc/Al) ratio in 

the soil solution, and the health of sugar maple and red spruce in forest ecosystems served as the 

biological indicator of the impacts of critical acid load exceedance. A main distinction between 

these two critical loads is that aquatic critical loads are largely an integrated function of the 

chemistry of run-off waters that feed the lake or stream within a watershed, while terrestrial 

critical acid loads are determined by the rooting zone section of the soil profile in a forest 

ecosystem. Therefore, it is possible to have different critical load values for aquatic and 

terrestrial ecosystems within the same watershed. 

The goal of this Task was to determine the relative degree of protection offered by 

aquatic versus terrestrial critical acid loads within a landscape. Critical acid loads for lakes and 

streams within watersheds of the Adirondacks and Shenandoah Valley were compared against 

terrestrial critical loads calculated for same watersheds to determine which estimate had the 

lowest, most protective critical load for acidifying nitrogen and sulfur deposition. 

September, 2010 B-1 Draft – Do Not Quote or Cite 
1 Methods

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For the REA Report (US EPA, 2009), critical acid loads were determined for 169 lakes 

and 60 streams in the Adirondacks and Shenandoah Case Study Areas, respectively. These 

critical loads were calculated using four different ANCs, 0, 20, 50 and 100 μeq/L, that ranged in 

the level of protection offered to fish species abundance and diversity, and the resulting critical 

acid loads were classified into four “current condition of acidity and sensitivity to acidification” 

categories. “Highly Sensitive” water bodies had critical loads less than or equal to 50 

meq/m2

/yr, “Moderately Sensitive” systems had critical loads ranging from 51 to 100 meq/m2/yr, 

“Low Sensitivity” lakes and streams had critical loads that ranged from 101 to 200 meq/m2

/yr, 

and “Not Sensitive” systems had critical acid loads greater than 201 meq/m2/yr. 

For the purposes of this Task, aquatic critical acid loads corresponding to an ANC of 50 

meq/m2

/yr were selected, and the locations of the lakes and streams in the Adirondacks and 

Shenandoah Case Study Areas were mapped by HUC12 watersheds. Availability of data for 

terrestrial acidification estimates was determined for each HUC, and only HUCs that had 

sufficient data were mapped. Data from the U.S. Department of Agriculture- Natural Resources 

Conservation Service (USDA-NRCS) SSURGO soils database (USDA-NRCS, 2008) had the 

poorest coverage. This data restriction limited the number of water bodies that could be included 

in the analysis to 62 and 35 for the Adirondacks and Shenandoah Case Study Areas, respectively. 

To examine a representative selection of water bodies in each Case Study Area, four 

watersheds containing lakes or streams from each of the four “current condition of acidity and 

sensitivity to acidification” categories were randomly selected. Therefore, a total of 16 

watersheds were chosen for each Case Study Area. All four “current condition of acidity and 

sensitivity to acidification” categories were evenly represented for the Adirondacks Case Study 

Area (four watersheds for each of the four categories). However, due to the limited number of 

watersheds in the Shenandoah Area and a lower proportion of lakes with low sensitivities to 

acidifying nitrogen and sulfur deposition (“Low Sensitivity” and “Not Sensitive”), it was not 

possible to have equal representation of all “current condition of acidity and sensitivity to 

acidification” categories. Therefore, there was a larger representation of streams that were more 

sensitive to acidification (“Highly Sensitive” and “Moderately Sensitive”). All water bodies that 

September, 2010 B-2 Draft – Do Not Quote or Cite 
September, 2010 B-3 Draft – Do Not Quote or Cite 

1

2

3

were located in each of the selected HUCs were included in the analyses. In many cases, these 

water bodies ranged in sensitivity to acidification. In total, 29 lakes and 20 streams were 

analyzed in the Adirondacks and Shenandoah Case Study Areas, respectively (Table 1 and 2). 
Table 1. Watersheds (HUC 12) and fresh water lakes in the Adirondacks Case Study Area that were used in the 

aquatic and terrestrial critical acid loads. Lake IDs and associated aquatic critical acid loads (CL) in meq/m2/yr, 

50 μeq/L, are indicated in each cell and are from the REA REPORT (US EPA, 2009). 

123

CURRENT CONDITION OF ACIDITY AND 

SENSITIVITY TO ACIDIFICATION CATEGORY 

HUC 

Highly Sensitive (CL 

< 50 meq/m2/yr) 

Moderately Sensitive (CL 

= 51-100 meq/m2/yr) 

Low Sensitivity 

(CL = 101-200 

meq/m2/yr) 

(CL 

020100010103 NY534

020100040203 1A2-028O (CL = 106) NY310L (CL = 147) NY30020100080304 NY31

NY31

020100081602 NY50020200020101 NY013L (CL = 64) 020200020704 NY536L (CL = 69) 

020200040805 1A2-078O (CL = 33) NY292L (CL = 117) 

041501011001 NY029L (CL = 39) 041503020801 NY78

041503040102 NY284L (CL = 23) NY285L (CL = 42) 

041503040204 NY278L (CL = 57) 

041503050103 1A1-089O (CL = 43) 050215AO (CL = 74) NY793L (CL = 97) 041503050104 NY290L (CL = 30) NY289L (CL = 50) 

September, 2010 B-4 Draft – 
CURRENT CONDITION OF ACIDITY AND 

SENSITIVITY TO ACIDIFICATION CATEGORY 

HUC 

Low Sensitivity Highly Sensitive (CL Moderately Sensitive (CL 

(CL = 101-200 < 50 meq/m2/yr) = 51-100 meq/m2/yr) (CL 

meq/m2/yr) 

041503050302 NY008L (CL = 146) NY007L (CL = 165) 041503050407 

NY767L (CL = 51) NY529L (CL = 73) NY528L (CL = 82) NY769L (CL = 99) 

NY768L (CL = 114) 

041503050601 NY004L (CL = 168) 

1

September, 2010 B-5 Draft – 
September, 2010 B-6 Draft – 

1234

Table 2. Watersheds (HUC 12) and streams in the Shenandoah Case Study Area that were used in the compariso

terrestrial critical acid loads. Stream IDs and associated aquatic critical acid loads (CL) in meq/m2/yr, based on a

are indicated in each cell and are from the REA REPORT (US EPA, 2009). 

CURRENT CONDITION OF ACIDITY AND SENSITIVITY TO ACIDIFICATIO

HUC 

Highly Sensitive (CL 

< 50 meq/m2/yr) 

Moderately Sensitive (CL 

= 51-100 meq/m2/yr) 

Low Sensitivity (CL 

= 101-200 meq/m2/yr) 

(C

020700050401 VT37 (CL = 26) 020700050502 VT57 (CL = 39) 020700050703 VT40 (CL = 13) 020700050705 VT35 (CL = 37) VT36 (CL = 24) 020700050801 DR01 (CL = 33) WOR1 (CL = 43) 020700050803 VT53 (CL = 40) 020700060101 VT54 (CL = 69) 020801030301 VT60 (CL = 198) VT020801030402 VT62 (CL = 68) 020802010702 VT10 (CL = 15) 020802010703 VT11 (CL = 14) VT12 (C = 75) 020802010801 VT14 (CL = 14) VT15 (CL = 13) 020802010803 VT16 (CL = 20) 020802020102 VT38 (CL = 66) 020802020401 VT41 (CL = 15) 020802030601 VT46 (CL = 52) 
Terrestrial critical acid loads were calculated for each of the 16 watersheds using the 

simple mass balance method (UNECE, 2004) and data sources outlined in the REA Report (US 

EPA, 2009), and Bc/Al soil solution indicator values of 1.2 and 10.0. Briefly, average values for 

base cation deposition (calcium, potassium, magnesium and sodium), chloride deposition, and 

annual runoff (m3/ha/yr) were determined for each watershed (Table 3). The Kgibb constant 

(m6

eq

2

) was determined by the average percent organic matter in the soil (Table 4), and N 

immobilization in the soil was set to the constant value of 42.86 eq/ha/yr (McNulty et al., 2007). 

It was assumed that active harvesting did not occur in each of the watersheds. Therefore base 

cation (calcium, magnesium and potassium) and nitrogen uptake were 0 eq/ha/yr (UNECE, 

2004). Similarly, it was assumed that the majority of each watershed consisted of upland sites. 

Therefore, denitrification losses were assumed to be 0 eq/ha/yr (McNulty et al., 2007). Base 

cation weathering was estimated using the clay substrate model (equations 1-3) (McNulty et al., 

2007). 

1

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8

9

10

11

12

13

 Acid Substrate:       2 14 e  clay%32.0clay%7.56BC (1) 

 Intermediate Substrate:       2 15 e  clay%18.0clay%6.53500BC (2) 

 Basic Substrate:  clay%2.59500BC  16 e    (3) 

17

18

19

20

21

22

23

24

25

26

27

28

29

30

where 

 BCe = empirical soil base cation (Ca2+ + K+ + Mg2+ + Na+) weathering rate 

(eq/ha/yr) 

 % clay = the percentage of clay within the top 50cm of the soil. 

The U.S. Department of Agriculture- Natural Resources Conservation Service (USDA-NRCS) 

SSURGO soils database (USDA-NRCS, 2008)) and state-level geology (U.S. Geological Survey 

(USGS) state-level integrated map database for the United States (USGS, 2009)) were used to 

determine parent material acidity classification. Parent material acidity was determined for each 

SSURGO polygon within each watershed using the criteria outlined in the REA Report (US 

EPA, 2009), and the contributions of base cations from the weathering of acid, intermediate and 

basic substrates (eq/ha/yr) were determined by a weighted average based on the proportion of 

area occupied by each parent material acidity class. Rooting depth was assumed to be 50 cm and 

masses of calcium, magnesium, potassium, sodium and nitrogen were converted to eq/ha/yr units 

September, 2010 B-7 Draft – Do Not Quote or Cite 
based on molar charge equivalents. Unless indicated otherwise, the units used in the calculation 

of critical acid loads were eq/ha/yr. The estimated terrestrial critical loads for the 16 watersheds 

in the Adirondacks and Shenandoah Case Study Areas are presented in Table 5. 

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7

8

Table 3. Name, type and source of data used in the simple mass balance estimates of terrestrial 

critical acid loads for the watersheds in the Adirondacks and Shenandoah Case Study Areas. 

DATA NAME TYPE SOURCE 

Base cation (Ca2+, 

Mg2+, Na+, K+) 

deposition— wet 

CMAQ/ 

NADP 

GIS 

datalayers 

Provided by U.S. Environmental 

Protection Agency (EPA)/NADP, 

2003a,c, d, e 

Chloride (Cl-) 

deposition— wet 

NADP GIS 

datalayer 

NADP, 2003b 

Runoff Annual run-off 

(1: 7,500,000 

scale) 

GIS 

datalayer 

Gebert et al., 1987 

Soil horizon depth SSURGO GIS 

datalayer 

USDA-NRCS, 2008 

Percentage of clay 

by soil horizon 

SSURGO GIS 

datalayer 

USDA-NRCS, 2008 

Percentage of 

organic matter by 

soil horizon 

SSURGO GIS 

datalayer 

USDA-NRCS, 2008 

Soil parent 

material 

SSURGO GIS 

datalayer 

USDA-NRCS, 2008 

State-level 

bedrock geology 

State 

Geological 

Map 

Compilation 

GIS 

datalayer 

USGS, 2009 

Note: CMAQ = Community Multiscale Air Quality Model; NADP = National Atmospheric 

Deposition Program; GIS = Geographic Information System; SSURGO = Soil Survey 

Geographic Database 

9

10

September, 2010 B-8 Draft – Do Not Quote or Cite 
Table 4. Gibbsite equilibrium (Kgibb) constant determined by percentage of soil organic matter 

(modified from McNulty et al. 2007). 

1

2

Soil Type Layer 

Organic 

Matter % Kgibb (m6/eq2) 

Mineral soils: C layer <5 950 

Soils with low organic matter: B/C layers 5 to15 300 

Soils with some organic material: A/E 

layers 15 to 70 100 

Peaty and organic soils: organic layers >70 9.5 

3

September, 2010 B-9 Draft – Do Not Quote or Cite 
Table 5. Terrestrial critical acid loads (in eq/ha/yr) for the watersheds in the Adirondacks and 

Shenandoah Case Study Areas. 

1

2

3

Terrestrial Critical Acid Load 

(eq/ha/yr) Case Study 

Area 

HUC12 

Bc/Al = 1.2 Bc/Al = 10.0 

Adirondacks 020100010103 2045 1134 

Adirondacks 020100040203 1316 712 

Adirondacks 020100080304 1329 731 

Adirondacks 020100081602 1670 922 

Adirondacks 020200020101 1484 819 

Adirondacks 020200020704 1707 935 

Adirondacks 020200040805 1770 951 

Adirondacks 041501011001 1770 955 

Adirondacks 041503020801 1664 912 

Adirondacks 041503040102 1627 880 

Adirondacks 041503040204 1436 786 

Adirondacks 041503050103 1774 957 

Adirondacks 041503050104 1794 968 

Adirondacks 041503050302 1754 947 

Adirondacks 041503050407 1447 789 

Adirondacks 041503050601 1203 656 

Shenandoah 020700050401 1440 802 

Shenandoah 020700050502 1560 871 

Shenandoah 020700050703 1762 979 

Shenandoah 020700050705 1852 1032 

Shenandoah 020700050801 1799 1003 

Shenandoah 020700050803 1975 1102 

Shenandoah 020700060101 1638 914 

Shenandoah 020801030301 1511 843 

Shenandoah 020801030402 1393 776 

Shenandoah 020802010702 1603 890 

Shenandoah 020802010703 1642 912 

Shenandoah 020802010801 1635 909 

Shenandoah 020802010803 1573 876 

Shenandoah 020802020102 1519 845 

Shenandoah 020802020401 1264 703 

Shenandoah 020802030601 1660 918 

September, 2010 B-10 Draft – Do Not Quote or Cite 
Maps were generated to compare the aquatic and terrestrial critical acid loads in each 

watershed to determine which estimate provided the greatest protection against acidifying 

nitrogen and sulfur deposition. In each watershed, the terrestrial critical load estimate was 

compared against each aquatic critical load, and the load with the lowest value was set to 

represent the most sensitive component in the watershed. All critical load estimates were 

converted to eq/ha/yr for the comparisons. 

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8 Results

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29

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31

Maps indicating and comparing the sensitivities of the terrestrial and aquatic critical loads 

to nitrogen and sulfur deposition in each watershed of the Adirondacks and Shenandoah Case 

Study Areas are presented in Figures 1-4 and Tables 6-9. 

In the Adirondacks Case Study Area, 7 of the 16 watersheds had terrestrial critical acid 

loads (based on a Bc/Al of 10.0) that were lower and therefore more sensitive to acidification 

than all the lakes in the watershed. However, when the terrestrial critical loads were calculated 

with a Bc/Al soil solution ratio of 1.2, only 5 of the 16 watersheds were protected by a terrestrial 

critical load that was lower than the aquatic critical loads of the lakes. Three watersheds in the 

Adriondacks Case Study Area had terrestrial critical loads (based on a Bc/Al of 10.0) that were 

lower and higher than the critical loads for the lakes in the watershed, and one watershed had a 

similar mixture of aquatic versus terrestrial acid load protections for terrestrial critical loads 

estimated with a Bc/Al of 1.2. In general, a main trend in the Adirondacks Case Study Area was 

that watersheds with “Highly Sensitive” and “Moderately Sensitive” lakes were more protected 

by aquatic than terrestrial critical acid loads, while the watersheds with “Low Sensitivity” and 

“Not Sensitive” lakes were more protected by terrestrial critical acid loads. 

Similar trends were found in the Shenandoah Case Study Area. However, there was little 

distinction between terrestrial acid loads that were calculated with a Bc/Al of 10.0 versus 1.2. 

Terrestrial critical acid loads offered a higher level of protection than did the stream aquatic 

critical loads in only one watershed. The two lakes in this watershed had “Low Sensitivity” or 

were “Not Sensitive” to acidifying nitrogen and sulfur deposition. The 15 watersheds that had 

streams with aquatic critical loads lower and more protective than the terrestrial critical loads, 

September, 2010 B-11 Draft – Do Not Quote or Cite 
were all “Highly Sensitive” or “Moderately Sensitive” to acidifying nitrogen and sulfur 

deposition. 

1

2

3

4

5

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8

9

10

11

12

13

14

15

 

In summary, a comparison of the terrestrial and aquatic critical acid loads for watersheds 

in the Adirondacks and Shenandoah Case Study Areas indicated that, in general, the aquatic 

critical acid loads offered greater protection to the watersheds than did the terrestrial critical 

loads. In situations where the terrestrial loads were more protective, the lakes or streams in the 

watershed were rated as having “Low Sensitivity” or “Not Sensitive” to acidifying nitrogen and 

sulfur deposition. Conversely, when the water bodies were more sensitive to deposition 

(“Highly Sensitive” or “Moderately Sensitive”), the aquatic critical acid loads consistently 

provided a greater level of protection against acidifying nitrogen and sulfur deposition in the 

watershed. 

September, 2010 B-12 Draft – Do Not Quote or Cite 
Figure 1. Comparison of aquatic and terrestrial critical loads of acidification (in eq/ha/yr) in the 

16 watersheds of the Adirondacks Case Study Area, based on an ANC of 50 eq/L for the aquatic 

loads and a Bc/Al of 10.0 for the terrestrial loads. Colored circles indicate the locations of the 

waters bodies within each watershed. Green circles indicate lakes with critical load values less 

than the terrestrial critical load for the same watershed. Red circles indicate a condition where 

the terrestrial critical load is lower than the lake critical load. 

1

2

3

4

5

6

7

8

September, 2010 B-13 Draft – Do Not Quote or Cite 
Figure 2. Comparison of aquatic and terrestrial critical loads of acidification (in eq/ha/yr) in the 

16 watersheds of the Adirondacks Case Study Area, based on an ANC of 50 eq/L for the aquatic 

loads and a Bc/Al of 1.2 for the terrestrial loads. Colored circles indicate the locations of the 

waters bodies within each watershed. Green circles indicate lakes with critical load values less 

than the terrestrial critical load for the same watershed. Red circles indicate a condition where 

the terrestrial critical load is lower than the lake critical load. 

1

2

3

4

5

6

7

8

9

September, 2010 B-14 Draft – Do Not Quote or Cite 
Figure 3. Comparison of aquatic and terrestrial critical loads of acidification (in eq/ha/yr) in the 

16 watersheds of the Shenandoah Case Study Area, based on an ANC of 50 eq/L for the aquatic 

loads and a Bc/Al of 10.0 for the terrestrial loads. Colored circles indicate the locations of the 

waters bodies within each watershed. Green circles indicate streams with critical load values 

less than the terrestrial critical load for the same watershed. Red circles indicate a condition 

where the terrestrial critical load is lower than the stream critical load. 

1

2

3

4

5

6

7

8

September, 2010 B-15 Draft – Do Not Quote or Cite 
September, 2010 B-16 Draft – Do Not Quote or Cite 

1

2

3

4

5

6

7

Figure 4. Comparison of aquatic and terrestrial critical loads of acidification (in eq/ha/yr) in the 

16 watersheds of the Shenandoah Case Study Area, based on an ANC of 50 eq/L for the aquatic 

loads and a Bc/Al of 1.2 for the terrestrial loads. Colored circles indicate the locations of the 

waters bodies within each watershed. Green circles indicate streams with critical load values 

less than the terrestrial critical load for the same watershed. Red circles indicate a condition 

where the terrestrial critical load is lower than the stream critical load. 

8
Table 6. Relative sensitivities of aquatic versus terrestrial critical loads in the 29 lakes and 16 watersheds of the 

Study Area (based on an ANC of 50 μeq/L for the aquatic loads and a Bc/Al of 10.0 for the terrestrial critical loaof eq/ha/yr) to acidifying nitrogen and sulfur deposition. Lake IDs are indicated in each cell and are from the REEPA, 2009). Green text indicates lakes where the aquatic critical load was less than the terrestrial critical load vawatershed. Red text indicates a condition where the terrestrial critical load for the watershed was lower than the for the lake within the same watershed. 

CURRENT CONDITION OF ACIDITY AND 

SENSITIVITY TO ACIDIFICATION CATEGORY 

HUC 

Highly Sensitive 

(CL < 50 

meq/m2/yr) 

Moderately Sensitive 

(CL = 51-100 

meq/m2/yr) 

Low Sensitivity

(CL = 101-200 

meq/m2/yr) 

Not 

(C

me

020100010103 N020100040203 1A2-028O NY310L N020100080304 N

N

020100081602 N020200020101 NY013L 020200020704 NY536L 020200040805 1A2-078O NY292L 041501011001 NY029L 041503020801 N041503040102 NY284L NY285L 041503040204 NY278L 041503050103 1A1-089O 050215AO NY793L041503050104 NY290L 

September, 2010 B-17 Draft – 
CURRENT CONDITION OF ACIDITY AND 

SENSITIVITY TO ACIDIFICATION CATEGORY 

HUC 

Highly Sensitive 

(CL < 50 

meq/m2/yr) 

Moderately Sensitive 

(CL = 51-100 

meq/m2/yr) 

Low Sensitivit

y Not 

(CL = 101-200 (C

meq/m2/yr) me

NY289L 

041503050302 NY008L NY007L 041503050407 

NY767L NY529L NY528L NY769L

NY768L 

041503050601 NY004L 

September, 2010 B-18 Draft – 
Table 7. Relative sensitivities of aquatic versus terrestrial critical loads in the 29 lakes and 16 watersheds of the 

Study Area (based on an ANC of 50 μeq/L for the aquatic loads and a Bc/Al of 1.2 for the terrestrial critical loadeq/ha/yr) to acidifying nitrogen and sulfur deposition. Lake IDs are indicated in each cell and are from the REA 

2009). Green text indicates lakes where the aquatic critical load was less than the terrestrial critical load value fo

text indicates a condition where the terrestrial critical load for the watershed was lower than the aquatic critical lo

the same watershed. 

CURRENT CONDITION OF ACIDITY AND 

SENSITIVITY TO ACIDIFICATION CATEGORY 

HUC Highly Sensitive 

(CL < 50 

meq/m2/yr) 

Moderately Sensitive 

(CL = 51-100 

meq/m2/yr) 

Low Sensitivity

(CL = 101-200 

meq/m2/yr) 

Not 

(C

me

020100010103 N020100040203 1A2-028O NY310L N020100080304 N

N

020100081602 N020200020101 NY013L 020200020704 NY536L 020200040805 1A2-078O NY292L 041501011001 NY029L 041503020801 N041503040102 NY284L NY285L 041503040204 NY278L 041503050103 1A1-089O 050215AO NY793L 

September, 2010 B-19 Draft – 
CURRENT CONDITION OF ACIDITY AND 

SENSITIVITY TO ACIDIFICATION CATEGORY 

HUC Hi

ghl

y Sensitive 

(CL < 50 

meq/m2/yr) 

Moderately Sensitive 

(CL = 51-100 

meq/m2/yr) 

Low Sensitivit

y Not 

(CL = 101-200 (C

meq/m2/yr) me

041503050104 NY290L NY289L 041503050302 NY008L NY007L 041503050407 

NY767L NY529L NY528L NY769L 

NY768L 

041503050601 NY004L 

September, 2010 B-20 Draft – 
Table 8. Relative sensitivities of aquatic versus terrestrial critical loads in the 20 streams and 16 watersheds of th

Study Area (based on an ANC of 50 μeq/L for the aquatic loads and a Bc/Al of 10.0 for the terrestrial critical loa

of eq/ha/yr) to acidifying nitrogen and sulfur deposition. Stream IDs are indicated in each cell and are from the R

EPA, 2009). Green text indicates streams where the aquatic critical load was less than the terrestrial critical load 

watershed. Red text indicates a condition where the terrestrial critical load for the watershed was lower than the 

for the stream within the same watershed. 

CURRENT CONDITION OF ACIDITY AND 

SENSITIVITY TO ACIDIFICATION CATEGORY 

HUC Highly Sensitive 

(CL < 50 

meq/m2/yr) 

Moderately Sensitive 

(CL = 51-100 

meq/m2/yr) 

Low Sensitivity

(CL = 101-200 

meq/m2/yr) 

Not S

(CL 

meq

020700050401 VT37 020700050502 VT57 020700050703 VT40 020700050705 VT35 VT36 020700050801 DR01 WOR1 020700050803 VT53 020700060101 VT54 020801030301 VT60 V020801030402 VT62 020802010702 VT10 020802010703 VT11 VT12 020802010801 VT14 VT15 

September, 2010 B-21 Draft – 
CURRENT CONDITION OF ACIDITY AND 

SENSITIVITY TO ACIDIFICATION CATEGORY 

HUC Highly Sensitive 

(CL < 50 

meq/m2/yr) 

Moderately Sensitive 

(CL = 51-100 

meq/m2/yr) 

Low Sensitivit

y Not S

(CL = 101-200 (CL 

meq/m2/yr) meq

020802010803 VT16 020802020102 VT38 020802020401 VT41 020802030601 VT46 

September, 2010 B-22 Draft – 
Table 9. Relative sensitivities of aquatic versus terrestrial critical loads in the 20 streams and 16 watersheds of th

Study Area (based on an ANC of 50 μeq/L for the aquatic loads and a Bc/Al of 1.2 for the terrestrial critical load

eq/ha/yr) to acidifying nitrogen and sulfur deposition. Stream IDs are indicated in each cell and are from the RE

EPA, 2009). Green text indicates streams where aquatic critical loads were less than the terrestrial critical load vwatershed. Red text indicates a condition where the terrestrial critical load for the watershed was lower than the 

for the stream within the same watershed.

CURRENT CONDITION OF ACIDITY AND 

SENSITIVITY TO ACIDIFICATION CATEGORY 

HUC Highly Sensitive 

(CL < 50 

meq/m2/yr) 

Moderately Sensitive 

(CL = 51-100 

meq/m2/yr) 

Low Sensitivity

(CL = 101-200 

meq/m2/yr) 

Not S

(CL 

meq

020700050401 VT37 020700050502 VT57 020700050703 VT40 020700050705 VT35 VT36 020700050801 DR01 WOR1 020700050803 VT53 020700060101 VT54 020801030301 VT60 V020801030402 VT62 020802010702 VT10 020802010703 VT11 VT12 020802010801 VT14 VT15 

September, 2010 B-23 Draft – 
September, 2010 B-24 Draft – 

CURRENT CONDITION OF ACIDITY AND 

SENSITIVITY TO ACIDIFICATION CATEGORY 

HUC Highly Sensitive 

(CL < 50 

meq/m2/yr) 

Moderately Sensitive 

(CL = 51-100 

meq/m2/yr) 

Low Sensitivity

(CL = 101-200 

meq/m2/yr) 

Not S

(CL 

meq

020802010803 VT16 020802020102 VT38 020802020401 VT41 020802030601 VT46 
REFERENCES 

Gebert, W.A., D.J. Graczyk, and W.R. Krug, 1987. Average Annual Runoff in the United States, 

1951-80: U.S. Geological Survey Hydrologic Investigations Atlas HA-710, Scale 

1:7,500,000. GIS datalayer. U.S. Department of Interior, U.S. Geological Survey, 

Madison, WI. Available at: http://water.usgs.gov/GIS/dsdl/runoff.e00.gz (accessed 

September 9, 2009). 

McNulty, S.G., E.C. Cohen, H. Li, and J.A. Moore-Myers. 2007. Estimates of critical acid loads 

and exceedences for forest soils across the conterminous United States. Environmental 

Pollution 149:281–292. 

NADP (National Atmospheric Deposition Program). 2003a. Annual Calcium Wet Deposition, 

2002. GIS datalayer. National Atmospheric Deposition Program, Illinois State Water 

Survey, Champaign, IL. Available at http://nadp.sws.uiuc.edu/maps/2002. 

NADP (National Atmospheric Deposition Program). 2003b. Annual Chloride Wet Deposition, 

2002. GIS datalayer. National Atmospheric Deposition Program, Illinois State Water 

Survey, Champaign, IL. Available at http://nadp.sws.uiuc.edu/maps/2002. 

NADP (National Atmospheric Deposition Program). 2003c. Annual Magnesium Wet Deposition, 

2002. GIS datalayer. National Atmospheric Deposition Program, Illinois State Water 

Survey, Champaign, IL. Available at http://nadp.sws.uiuc.edu/maps/2002. 

NADP (National Atmospheric Deposition Program). 2003d. Annual Potassium Wet Deposition, 

2002. GIS datalayer. National Atmospheric Deposition Program, Illinois State Water 

Survey, Champaign, IL. Available at http://nadp.sws.uiuc.edu/maps/2002. 

NADP (National Atmospheric Deposition Program). 2003e. Annual Sodium Wet Deposition, 

2002. GIS datalayer. National Atmospheric Deposition Program, Illinois State Water 

UNECE (United Nations Economic Commission for Europe). 2004. Manual on Methodologies 

and Criteria for Modeling and Mapping Critical Loads and Levels and Air Pollution 

Effects, Risks, and Trends. Convention on Long-Range Transboundary Air Pollution, 

Geneva Switzerland. Available at http://www.icpmapping.org (accessed August 16, 

2006). 

USDA-NRCS (United States Department of Agriculture-Natural Resources Conservation 

Service). 2008. Soil Survey Geographic (SSURGO) Database. GIS datalayer. U.S. 

Department of Agriculture, Natural Resources Conservation Service, Washington, DC. 

Available at http://datagateway.nrcs.usda.gov. 

US EPA (United States Environmental Protection Agency). 2009. Risk and Exposure Assessment 

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

Nitrogen and Sulfur. Final. U.S. Environmental Protection Agency, Office of Research 

and Development, National Center for Environmental Assessment, Research Triangle 

Park, NC. September. 

USGS (U.S. Geological Survey). 2009. State Geological Map Compilation. U.S. Department of 

the Interior, U.S. Geological Survey, Reston, VA. Available at: 

http://tin.er.usgs.gov/geology/state (accessed January 28, 2009). 

September, 2010 B-25 Draft – Do Not Quote or Cite 
1

2

3

4

5

6

APPENDIX C: Elasticity AAPI sensitivity analyses 

Elasticity. One metric for determining sensitivity of the AAPI to its component parameters is 

elasticity. Elasticity measures the percent change in AAPI for a one percent change in the 

component. In general, the formula for an elasticity is: 

AAPI

X

X

AAPI E j

j

AAPI

X j  



7 

8

9

10

11

12

Where is the elasticity of AAPI with respect to component Xj, and j is the number of components. AAPI

X j E

So, for AAPI defined as

 SOxTNOyT  Q

NHx

Q

BCN

Q

AAPI eco NOy SOx 

1 * 11

0 13

14

15

16

17

18

The set of relevant elasticities are:

For runoff, Q:

  AAPI

Q SOxTNOyTNHxN Q

E eco NOy SOx

AAPI

Q   2

1

19 , which can be rewritten as

20

  AAPI

Q BCAAPI

Q

E AAPI

Q  *0

1

21 , or

22

September, 2010 C-1 Draft – Do Not Quote or Cite 
AAPI

BC E AAPI

Q

*

0 1 1

2

3

4

For , * BC0

AAPI

BC E AAPI

Q

*

0 5 

6

7

8

For Neco,

AAPI

N

Q

E AAPI eco

Neco  1

9

10

11

12

For NHx,

AAPI

NHx

Q

E AAPI

NHx  1

13

14

15

16

For TNOy,

AAPI

T

NOy QE AAPI NOy

TNOy  1

17

18

19

20

21

For TSOx,

September, 2010 C-2 Draft – Do Not Quote or Cite 
AAPI

T SOx

Q

E AAPI SOx

TSOx  1

1

2

3

4

For NOy,

AAPI

NOy T

Q

E NOy

AAPI

NOy  1

5

6

7

8

For SOx,

AAPI

SOx T

Q

E SOx

AAPI

SOx  1

9

10

11

12

13

14

15

16

17

18

19

20

21

22

23

24

25

26

27

28

29

These elasticities can be evaluated at various points along the ranges of each component, as well as along 

ranges of the AAPI. We evaluate the elasticities at the sample means, medians, first quartiles, and third 

quartiles. Elasticities are evaluated only for ecoregions that overlap the CMAQ modeling domain which 

provides values for reduced nitrogen and the transformation ratios (TNOx and TSOx). This will provide a 

reasonable assessment of the sensitivity of the AAPI to input components. Table 1 provides the estimated 

elasticities. Elasticities are summarized across ecoregions using means, medians, minimums, and 

maximums. 

Note that elasticities can be either positive or negative. A negative elasticity means that the calculated 

AAPI will decrease as a component increases. The magnitude of the elasticity depends on the values of 

the components and the starting value of AAPI. 

Based on the calculated elasticities, AAPI is most responsive to changes in Q, BC0, and Neco with some 

responsiveness to reduced N. Note that for some components, such as Q, the elasticities switch signs 

depending on the values of the variables for which the elasticity is evaluated. This suggests potentially 

important interactions. AAPI is not responsive to the transformation ratios, TNOx and TSOx at mean 

values of the AAPI components. However, when the elasticities for TNOx and TSOx are evaluated at the 

first quartiles of the data, some locations in the Eastern U.S. show higher responsiveness to changes in 

TNOx and TSOx, with elasticities as high as 2. 

September, 2010 C-3 Draft – Do Not Quote or Cite 
1 Table 1. Elasticity* of AAPI to Component Variables 

AAPI 

Component 

Metric for 

Which 

Elasticities 

are 

Evaluated 

Mean 

Elasticity 

Across 

Ecoregions 

Median 

Elasticity 

Across 

Ecoregions 

Minimum 

Elasticity 

Across 

Ecoregions 

Maximum 

Elasticity 

Across 

Ecoregions 

Mean -0.1047 0.1221 -20.4572 1.6005

Median 0.2561 0.1426 -6.2481 2.4578

1st Quartile -0.9283 0.1303 -135.2544 88.1063

Runoff (Q) 

3rd Quartile 0.2988 0.1110 -0.1810 7.5684

Mean 0.8953 1.1221 -19.4572 2.6005

Median 1.2561 1.1426 -5.2481 3.4578

1st Quartile 0.0717 1.1303 -134.2544 89.1063

Base Cation Weathering (BC0) 

3rd Quartile 1.2988 1.1110 0.8190 8.5684

Mean 0.0179 0.1464 -13.8545 1.6051

Median 0.3376 0.2563 -4.7203 2.5044

1st Quartile 0.3543 0.2596 -115.1112 137.0526

Neco 

3rd Quartile 0.3016 0.1440 0.0137 6.5565

Mean -0.0409 -0.0702 -0.4708 3.9332

Median -0.1407 -0.1031 -0.7957 1.0061

1st Quartile -0.0308 -0.1190 -33.9063 35.3783

Reduced Nitrogen 

3rd Quartile -0.1128 -0.0615 -1.9167 -0.0050

Mean 0.0089 -0.0053 -0.0597 1.0598

Median -0.0061 -0.0064 -0.0506 0.2751

NOx Transformation Ratio (TNOx) 

1st Quartile -0.0191 -0.0071 -2.7061 1.7749

3rd Quartile ‐0.0154 ‐0.0044 ‐0.5028 ‐0.0002

September, 2010 C-4 Draft – Do Not Quote or Cite 
September, 2010 C-5 Draft – Do Not Quote or Cite 

AAPI 

Component 

Metric for 

Which 

Elasticities 

are 

Evaluated 

Mean 

Elasticity 

Across 

Ecoregions 

Median 

Elasticity 

Across 

Ecoregions 

Minimum 

Elasticity 

Across 

Ecoregions 

Maximum 

Elasticity 

Across 

Ecoregions 

Mean 0.0019 ‐0.0024 ‐0.0185 0.3608

Median ‐0.0045 ‐0.0036 ‐0.0413 0.1091

1st Quartile 0.0040 ‐0.0032 ‐1.8023 1.9860

SOx TransformationRatio (TSOx)

3rd Quartile ‐0.0058 ‐0.0021 ‐0.1534 ‐0.0002

1

2

3

4

5

*Elasticity is the percent change in AAPI for a one percent change in the component variable. 

For example, when evaluated at the means of all component variables, the mean elasticity of 

AAPI to the runoff variable Q is -0.1047, which means that for each 1 percent increase in Q, the 

AAPI is reduced by 0.1047 percent. 
United States 

Environmental Protection Agency 

Office of Air Quality Planning and Standards 

Health and Environmental Impacts Division 

Research Triangle Park, NC 

Publication N