(4) Based on the steady-state critical load model for the year 2002, 52%, 72%, 85%, and 93% of 60 modeled streams received combined total sulfur and nitrogen deposition that exceeded their critical load, with critical ANC limits of 0, 20, 50, and 100 μeq/L respectively.
(5) Based on a deposition scenario that maintains current emission levels to 2020 and 2050, the simulation forecast indicates that a large number of streams still have Elevated to Acute problems with acidity. In fact, from 2006 to 2050, the percentage of streams with Acute Concern increases by 5%, while the percentage of streams in Moderate Concern decreases by 5%.
a. Deposition-related terrestrial acidification
	The role of terrestrial acidification was examined in the REA using a critical load analysis for sugar maple and red spruce forests in the eastern United States by using the base cation to aluminum (Bc/Al) ratio in acidified forest soils as an indicator to assess the impact of nitrogen and sulfur deposition on tree health. These are the two most commonly studied species in North America for impacts of acidification. At a Bc/Al ratio of 1.2, red spruce growth can be reduced by 20%. Sugar maple growth can be reduced by 20% at a Bc/Al ratio of 0.6.  Key findings of the case study are summarized below.
(1) Case study results suggest that the health of at least a portion of the sugar maple and red spruce growing in the United States may have been compromised with acidifying total nitrogen and sulfur deposition in 2002.  2002 CMAQ/NADP total nitrogen and sulfur deposition levels exceeded three selected critical loads in 3% to 75% of all sugar maple plots across 24 states. The three critical loads ranged from 107 to 6,008 eq/ha/yr for the Bc/Al ratios of 0.6, 1.2, and 10.0 (increasing levels of tree protection). The 2002 CMAQ/NADP total nitrogen and sulfur deposition levels exceeded three selected critical loads in 3% to 36% of all red spruce plots across 8 states. The three critical loads ranged from 180 to 4,278 eq/ha/yr for the Bc/Al ratios of 0.6, 1.2, and 10.0 (increasing levels of tree protection).
(2) The Simple Mass Balance model assumptions made for base cation weathering (Bcw) and forest soil ANC input parameters are the main sources of uncertainty since these parameters are rarely measured and require researchers to use default values. Bcw contributed 49% to the total variability in the critical load estimates, and forest soil ANC contributed 46% to the total variability.
(3) The pattern of case study results suggests that nitrogen and sulfur acidifying deposition in the sugar maple and red spruce forest areas studied were very close to, if not greater than, the critical loads for those areas and both ecosystems are likely to be sensitive to any future changes in the levels of deposition.
a. Deposition-related aquatic nutrient enrichment
	The role of nitrogen deposition in two main stem rivers feeding their respective estuaries was analyzed in the REA to determine if decreases in deposition could influence the risk of eutrophication as predicted using the Assessment of Estuarine Trophic Status eutrophication index (ASSETS EI) scoring system in tandem with SPAtially Referenced Regression on Watershed Attributes (SPARROW) modeling. This modeling approach provides a transferrable, intermediate-level analysis of the linkages between atmospheric deposition and receiving waters, while providing results on which conclusions could be drawn. Future application of the methods to case study areas where atmospheric deposition plays a larger role in the nitrogen loading to an estuary will likely provide more tangible results. A summary of findings follows:
(1) 2002 CMAQ/NADP results showed that an estimated 40,770,000 kg of total nitrogen was deposited in the Potomac River watershed. SPARROW modeling predicted that 7,380,000 kg N/yr of the deposited nitrogen reached the estuary (20% of the total load to the estuary). The overall ASSETS EI for the Potomac River and Potomac Estuary was Bad.
(2) To improve the Potomac River and Potomac Estuary ASSETS EI score from Bad to Poor, there is a slim chance that a decrease of at least 78% in the 2002 total nitrogen atmospheric deposition load to the watershed would be required.
(3) 2002 CMAQ/NADP results showed that an estimated 18,340,000 kg of total nitrogen was deposited in the Neuse River watershed. SPARROW modeling predicted that 1,150,000 kg N/yr of the deposited nitrogen reached the estuary (26% of the total load to the estuary). The overall ASSETS EI for the Neuse River/Neuse River Estuary was Bad.
(4) It was found that the Neuse River/Neuse River Estuary ASSETS EI score could not be improved from Bad to Poor with decreases only in the 2002 atmospheric deposition load to the watershed. Additional reductions would be required from other nitrogen sources within the watershed.
	The small effect of decreasing atmospheric deposition in the Neuse River watershed is because the other nitrogen sources within the watershed are more influential than atmospheric deposition to the total nitrogen loadings to the Neuse River Estuary as estimated with the SPARROW model. A waterbody's response to nutrient loading depends on the magnitude (e.g., agricultural sources have a high influence in the Neuse), spatial distribution, and other characteristics of the sources within the watershed.
e. Deposition-related terrestrial nutrient enrichment
	California Coastal Sage Scrub (CSS) and Mixed Conifer Forest (MCF) communities were the focus of the Terrestrial Nutrient Enrichment Case Studies of the REA. 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. These benchmarks, ranging from 3.1 to 17 kg N/ha/yr, were compared to 2002 CMAQ/NADP data to discern any associations between atmospheric deposition and changing communities. Evidence supports the finding that nitrogen alters CSS and MCF. Key findings include the following:
(1) 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.
(2) 2002 CMAQ/NADP nitrogen deposition data exceeds the 3.1 kg N/ha/yr benchmark in more than 38% (1,099,133 ha) of MCF areas, and nitrate leaching has been observed in surface waters. Ozone effects confound nitrogen effects on MCF acidophyte lichen, and the interrelationship between fire and nitrogen cycling requires additional research.
f. 	Additional effects
	Ecological effects have also been documented across the United States where elevated nitrogen deposition has been observed, including the eastern slope of the Rocky Mountains where shifts in dominant algal species in alpine lakes have occurred where wet nitrogen deposition was only about 1.5 kg N/ha/yr. High alpine terrestrial communities have a low capacity to sequester nitrogen deposition, and monitored deposition exceeding 3 to 4 kg N/ha/yr could lead to community-level changes in plant species, lichens, and mycorrhizae.
	Additional welfare effects are documented, but examined less extensively, in the REA.  These effects include qualitative discussions related to visibility and materials damage, such as corrosion, erosion, and soiling of paint and buildings which are being addressed in the particulate matter (PM) NAAQS review currently underway.  A discussion of the causal relationship between sulfur deposition (as sulfate, SO4[2-]) and increased mercury methylation in wetlands and aquatic environments is also included in the REA. On this subject the REA concludes that decreases in sulfate deposition will likely result in decreases in methyl mercury concentration; however, spatial and biogeochemical variations nationally hinder establishing large scale dose-response relationships.
	Several additional issues concerning oxides of nitrogen were addressed in the REA.  Consideration was also given to nitrous oxide (N2O), a potent GHG.  The REA concluded that it is most appropriate to analyze the role of N2O in the context of all of the GHGs rather than as part of the REA for this review.  The REA considered nitrogen deposition and its correlation with the rate of photosynthesis and net primary productivity. Nitrogen addition ranging from 15.4 to 300 kg N/ha/yr is documented as increasing wetland N2O production by an average of 207%. Nitrogen addition ranging from 30 to 240 kg N/ha/y increased CH4 emissions by 115%, averaged across all ecosystems, and methane uptake was reduced by 38% when nitrogen addition ranged from 10 to 560 kg N/ha/yr, but reductions were only significant for coniferous and deciduous forests. The heterogeneity of ecosystems across the United States, however, introduces variations into dose-response relationships.
	The phytotoxic effects of oxides of nitrogen and sulfur on vegetation were also briefly discussed in the REA which concluded that since a unique secondary NAAQS exists for SO2, and concentrations of nitric oxide (NO), NO2, and PAN are rarely high enough to have phytotoxic effects on vegetation that further assessment was not warranted at this time.
2. Conclusions on Effects
	For aquatic and terrestrial acidification effects, a similar conceptual approach was used (critical loads) to evaluate the impacts of multiple pollutants on an ecological endpoint, whereas the approaches used for aquatic and terrestrial nutrient enrichment were fundamentally distinct. Although the ecological indicators for aquatic and terrestrial acidification (i.e., ANC and Bc/Al) are very different, both ecological indicators are well-correlated with effects such as reduced biodiversity and growth. While aquatic acidification is clearly the targeted effect area with the highest level of confidence, the relationship between atmospheric deposition and an ecological indicator is also quite strong for terrestrial acidification. The main drawback with the understanding of terrestrial acidification is that the data are based on laboratory responses rather than field measurements. Other stressors that are present in the field but that are not present in the laboratory may confound this relationship.
	For nutrient enrichment effects, the REA utilized different types of indicators for aquatic and terrestrial effects to assess both the likelihood of adverse effects to ecosystems and the relationship between adverse effects and atmospheric sources of oxides of nitrogen.  The ecological indicator chosen for aquatic nutrient enrichment, the ASSETS EI, seems to be inadequate to relate atmospheric deposition to the targeted ecological effect, likely due to the many other confounding factors. Further, there is far less confidence associated with the understanding of aquatic nutrient enrichment because of the large contributions from non-atmospheric sources of nitrogen and the influence of both oxidized and reduced forms of nitrogen, particularly in large watersheds and coastal areas. However, a strong relationship exists between atmospheric deposition of nitrogen and ecological effects in high alpine lakes in the Rocky Mountains because atmospheric deposition is the only source of nitrogen to these systems. There is also a strong weight-of-evidence regarding the relationships between ecological effects attributable to terrestrial nitrogen nutrient enrichment; however, ozone and climate change may be confounding factors. In addition, the response for other species or species in other regions of the United States has not been quantified.
C. Adversity of Effects to Public Welfare
	Characterizing a known or anticipated adverse effect to public welfare is an important component of developing any secondary NAAQS. According to the 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 linking adversity to public welfare to disruptions in ecosystem structure and function has been used broadly by EPA to categorize effects of pollutants from the cellular to the ecosystem level.  An evaluation of adversity to public welfare might consider the likelihood, type, magnitude, and spatial scale of the effect as well as the potential for recovery and any uncertainties relating to these considerations.
	Similar concepts were used in past reviews of secondary NAAQS for ozone, PM (relating to visibility), as well as in initial reviews of effects from lead deposition.  Because oxides of nitrogen and sulfur are deposited from ambient sources into ecosystems where they affect changes to organisms, populations and ecosystems, the concept of adversity to public welfare as related to impacts on the public from alterations in structure and function of ecosystems is an appropriate consideration for this review.
	Based on information provided in the Policy Assessment, the following section discusses how ecological effects from deposition of oxides of nitrogen and sulfur relate to adversity to public welfare.  In the Policy Assessment, public welfare was discussed in terms of loss of ecosystem services (defined below), which in some cases can be monetized.  Each of the four main effect areas (aquatic and terrestrial acidification and aquatic and terrestrial nutrient over-enrichment) are discussed including current ecological effects and associated ecosystem services, including potential value of the services.
1. Ecosystem Services
	The Policy Assessment defines ecosystem services as the benefits individuals and organizations obtain from ecosystems.  Ecosystem services can be classified as provisioning (food and water), regulating (control of climate and disease), cultural (recreational, existence, spiritual, educational), and supporting (nutrient cycling).  Conceptually, changes in ecosystem services may be used to aid in characterizing a known or anticipated adverse effect to public welfare.  In the REA and Policy Assessment ecosystem services are discussed as a method of assessing the magnitude and significance to the public of resources affected by ambient concentrations of oxides of nitrogen and sulfur and deposition in sensitive ecosystems.
	As EPA has in previous NAAQS reviews defined ecological goods and services for the purposes of a Regulatory Impact Analysis as the "outputs of ecological functions or processes that directly or indirectly contribute to social welfare or have the potential to do so in the future.  Some outputs may be bought and sold, but most are not marketed."   It is especially important to acknowledge that it is difficult to measure and/or monetize the goods and services supplied by ecosystems.  It can be informative in characterizing adversity to public welfare to attempt to place an economic valuation on the set of goods and services that have been identified with respect to a change in policy however it must be noted that this valuation will be incomplete and illustrative only.
	Knowledge about the relationships linking ambient concentrations and ecosystem services is considered in the Policy Assessment as one method by which to inform a policy judgment on a known or anticipated adverse public welfare effect.  For example, a change in an ecosystem structure and process, such as foliar injury, would be classified as an ecological effect, with the associated changes in ecosystem services, such as primary productivity, food availability, forest products, and aesthetics (e.g., scenic viewing), classified as public welfare effects.  Additionally, changes in biodiversity would be classified as an ecological effect, and the associated changes in ecosystem services -- productivity, existence (nonuse) value, recreational viewing and aesthetics -- would also be classified as public welfare effects.
	As described in Chapters 4 and 5 of the REA, case study analyses were performed that link deposition in sensitive ecosystems to changes in a given ecological indicator (e.g., for aquatic acidification, to changes in acid neutralizing capacity [ANC]) and then to changes in ecosystems.  Appendix 8 of the REA links the changes in ecosystems to the services they provide (e.g., fish species richness and its influence on recreational fishing). To the extent possible for each targeted effect area, the REA linked ambient concentrations of nitrogen and sulfur (i.e., ambient air quality indicators) to deposition in sensitive ecosystems (i.e., exposure pathways), and then to system response as measured by a given ecological indicator (e.g., lake and stream acidification as measured by ANC). The ecological effect (e.g., changes in fish species richness) was then, where possible, associated with changes in ecosystem services and the corresponding public welfare effects (e.g., recreational fishing).
2. Effects on Ecosystem Services
	The process used to link ecological indicators to ecosystem services is discussed extensively in Appendix 8 of the REA.  In brief, for each case study area assessed the ecological indicators are linked to an ecological response that is subsequently linked to associated services to the extent possible.  For example in the case study for aquatic acidification the chosen ecological indicator is ANC which can be linked to the ecosystem service of recreational fishing.  Although recreational fishing losses are the only service effects that can be independently quantified or monetized at this time, there are numerous other ecosystem services that may be related to the ecological effects of acidification.
	While aquatic acidification is the focus of this proposed standard, the other effect areas were also analyzed in the REA and these ecosystems are being harmed by nitrogen and sulfur deposition and will obtain some measure of protection with any decrease in that deposition regardless of the reason for the decrease. The following summarizes the current levels of specific ecosystem services for aquatic and terrestrial acidification, and aquatic and terrestrial nutrient over-enrichment and attempts to quantify and when possible monetize the harm to public welfare, as represented by ecosystem services, due to nitrogen and sulfur deposition.
a. Aquatic Acidification
	Acidification of aquatic ecosystems primarily affects the ecosystem services that are derived from the fish and other aquatic life found in surface waters.  In the northeastern United States, the surface waters affected by acidification are not a major source of commercially raised or caught fish; however, they are a source of food for some recreational and subsistence fishers and for other consumers. Although data and models are available for examining the effects on recreational fishing, relatively little data are available for measuring the effects on subsistence and other consumers.  Inland waters also provide aesthetic and educational services along with non-use services, such as existence value (protection and preservation with no expectation of direct use).  In general, inland surface waters such as lakes, rivers, and streams also provide a number of regulating services, playing a role in hydrological regimes and climate regulation. There is little evidence that acidification of freshwaters in the northeastern 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.  Although it is difficult to quantify these services and how they are affected by acidification, it is worth noting that some of these services may be captured through measures of provisioning and cultural services. For example, these biological control services may serve as "intermediate" inputs that support the production of "final" recreational fishing and other cultural services.
	As summarized in Chapter 4 of the Policy Assessment, recent studies indicate that acidification of lakes and streams can result in significant loss in economic value.  For example, data indicate that more than 9% of adults in the northeastern part of the country participate annually in freshwater fishing yielding 140 million freshwater fishing days.  Each fishing day has an estimated average value per day of $35. Therefore, the implied total annual value of freshwater fishing in the northeastern United States was $5 billion in 2006.  Embedded in these numbers is a degree of harm to recreational fishing services due to acidification that has occurred over time.  These harms have not been quantified on a regional scale; however, a case study was conducted in the Adirondacks area (US EPA, 2011, section 4.4.2).
	In the Adirondacks case study, estimates of changes in recreational fishing services were determined, as well as changes more broadly in "cultural" ecosystem services (including recreational, aesthetic, and nonuse services).  First, the MAGIC model (US EPA, 2009, Appendix 8 and section 2.2)  was applied to 44 lakes to predict what ANC levels would be under both "business as usual" conditions (i.e., allowing for some decline in deposition due to existing regulations) and pre-emission (i.e., background) conditions.  Second, to estimate the recreational fishing impacts of aquatic acidification in these lakes, an existing model of recreational fishing demand and site choice was applied.  This model predicts how recreational fishing patterns in the Adirondacks would differ and how much higher the average annual value of recreational fishing services would be for New York residents if lake ANC levels corresponded to background (rather than business as usual) conditions.  To estimate impacts on a broader category of cultural (and some provisioning)  ecosystem services, results from the Banzhaf et al (2006) valuation survey of New York residents were adapted and applied to this context.  The survey used a contingent valuation approach to estimate the average annual household WTP for future reductions in the percent of Adirondack lakes impaired by acidification.  The focus of the survey was on impacts on aquatic resources. Pretesting of the survey indicated that respondents nonetheless tended to assume that benefits would occur in the condition of birds and forests as well as in recreational fishing.
	By extrapolating the 44 lake Adiriondack case study to all 3,000 Adirondack lakes and by applying the willingness to pay survey results to all New York residents, the study estimated aggregated benefits between $300 and $800 million annually for the equivalent of improving lakes in the Adirondacks region to an ANC level of 50 ueq/L.  The REA estimated 44% of the Adirondack lakes currently fall below an ANC of 50 ueq/L.  Several states have set goals for improving the acid status of lakes and streams, generally targeting ANC in the range of 50 to 60 ueq/L, and have engaged in costly activities to decrease acidification.
 	These results imply significant value to the public in addition to those derived from recreational fishing services. Note that the results are only applicable to improvements in the Adirondacks valued by residents of New York.  If similar benefits exist in other acid-impacted areas, benefits for the nation as a whole could be substantial. The analysis provides results on only a subset of the impacts of acidification on ecosystem services and suggests that the overall impact on these services is likely to be substantial.
b. Terrestrial Acidification
	Chapter 4.4.3 and 4.4.4 of the Policy Assessment review several economic studies of areas sensitive to terrestrial acidification.  Forests in the northeastern United States provide several important and valuable provisioning ecosystem services, which are reflected in the production and sales of tree products. Sugar maples are a particularly important commercial hardwood tree species in the United States, producing timber and maple syrup that provide hundreds of millions of dollars in economic value annually.   Red spruce is also used in a variety of wood products and provides up to $100 million in economic value annually.  Although the data do not exist to directly link acidification damages to economic values of lost recreational ecosystem services in forests, these resources are valuable to the public.  A recent study, reviewed in the Policy Assessment, suggests that the total annual value of recreational off-road driving was more than $9 billion and the value of hunting and wildlife viewing was more than $4 billion each in the northeastern United States.
c. Nutrient Enrichment
	Chapter 4.4.5 and 4.4.6 of the Policy Assessment summarize economic studies of east coast estuaries affected by nutrient over-enrichment or eutrophication.  Estuaries in the eastern United States are important for fish and shellfish production. The estuaries are capable of supporting large stocks of resident commercial species, and they serve as the breeding grounds and interim habitat for several migratory species. To provide an indication of the magnitude of provisioning services associated with coastal fisheries, from 2005 to 2007, the average value of total catch was $1.5 billion per year in 15 East Coast states.  Estuaries also provide an important and substantial variety of cultural ecosystem services, including water-based recreational and aesthetic services. For example, data indicate that 4.8% of the population in coastal states from North Carolina to Massachusetts participated in saltwater fishing, with a total of 26 million saltwater fishing days in 2006.  Based on estimates in the Policy Assessment, total recreational value from these saltwater fishing days was approximately $1.3 billion.  Recreational participation estimates for 1999 - 2000 showed almost 6 million individuals participated in motorboating in coastal states from North Carolina to Massachusetts. The aggregate value of these coastal motorboating outings was $2 billion per year.
	Terrestrial ecosystems can also suffer from nutrient over-enrichment.  Each ecosystem is different in its composition of species and nutrient requirements.  Changes to individual ecosystems from changes in nitrogen deposition can be hard to assess economically.  Relative recreational values are often determined by public use information.  Chapter 4.4.7 of the Policy Assessment reviewed studies related to park use in California.  Data from California State Parks 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 Policy Assessment indicate that the aggregate annual benefit for California residents from trail hiking in 2007 was $11.59 billion.
	The Policy Assessment also identified fire regulation as a service that could be affected by nutrient over-enrichment of the Coastal Sage Scrub and Mixed Conifer Forest ecosystems by encouraging growth of more flammable grasses, increasing fuel loads, and altering the fire cycle. Over the 5-year period from 2004 to 2008, Southern California experienced, on average, over 4,000 fires per year, burning, on average, over 400,000 acres per year.  It is not possible at this time to quantify the contribution of nitrogen deposition, among many other factors, to increased fire risk.
1. Summary
	Adversity to public welfare can be understood not only by looking at how deposition of oxides of nitrogen and sulfur affect the ecological functions of an ecosystem (see II.A.), but also by understanding the ecosystem services that are degraded.  The monetized value of the ecosystem services provided by ecosystems that are sensitive to deposition of oxides of nitrogen and sulfur are in the billions of dollars each year.  Many lakes and streams are known to be degraded by acidic deposition which affects recreational fishing and tourism.  Forest growth is likely suffering from acidic deposition in sensitive areas affecting red spruce and sugar maple timber production, sugar maple syrup production, hiking, aesthetic enjoyment and tourism.  Nitrogen deposition contributes significantly to eutrophication in many estuaries affecting fish production, swimming, boating, aesthetic enjoyment, and tourism.  Important biodiversity is likely affected in many natural and scenic terrestrial areas affecting habitat for rare and endangered species, fire control, hiking, aesthetic enjoyment, and tourism.
D.	Adequacy of the Current Standards
	An important issue to be addressed in the current review of the secondary standards for oxides and nitrogen and sulfur standard is whether, in view of the scientific evidence reflected in the ISA, additional information on exposure and risk discussed in the REA, and conclusions drawn from the Policy Assessment, the existing standards provide adequate protection.   The Administrator therefore, has considered the extent to which the current standards are adequate for the protection of public welfare.    Having reached this general conclusion, it is then necessary to first evaluate the appropriateness of the current standards to address the ecological effects of oxides of nitrogen and sulfur as well as the adequacy of the current secondary standards for oxides of nitrogen and sulfur to provide requisite protection by considering to what degree risks to sensitive ecosystems would be expected to occur in areas that meet the current standards.  Conclusions regarding the adequacy of the current standards are based on the available ecological effects, exposure and risk-based evidence.   In evaluating the strength of this information, EPA has taken into account the uncertainties and limitations in the scientific evidence.  This section addresses the adequacy of the current standards to protect against direct exposure effects on plants from oxides of nitrogen and sulfur, the appropriateness of the current structure of the standards to address deposition-related effects of oxides of nitrogen and sulfur on sensitive ecosystems and finally, the adequacy of such standards to protect against adverse effects related to the deposition of oxides of nitrogen and sulfur.
1.	Adequacy of the Current Standards for Direct Effects
	The current secondary oxides of nitrogen and sulfur standards are intended to protect against adverse effects to public welfare.  For oxides of nitrogen, the current secondary standard was set identical to the primary standard, e.g. an annual standard set for NO2 to protect against adverse effects on vegetation from direct exposure to ambient oxides of nitrogen.  For oxides of sulfur, the current secondary standard is a 3-hour standard intended to provide protection for plants from the direct foliar damage associated with atmospheric concentrations of SO2.  It is appropriate to consider whether the current standards are adequate to protect against the direct effects on vegetation resulting from ambient NO2 and SO2 which were the basis for the current secondary standards.  The ISA concluded that there was sufficient evidence to infer a causal relationship between exposure to SO2, NO, NO2 and PAN and injury to vegetation.  Additional research on acute foliar injury has been limited and there is no evidence to suggest foliar injury below the levels of the current secondary standards for oxides of nitrogen and sulfur.  There is sufficient evidence to suggest that the levels of the current standards are likely adequate to protect against direct phytotoxic effects.
1. Appropriateness and Adequacy of the Current Standards for Deposition-related Effects
	This section addresses two concepts necessary to evaluate the current standards on the context of deposition related effects.  First, appropriateness of the current standards is considered with regard to indicator, form, level and averaging time.  This discussion centers around the ability of the current standards to evaluate and provide protection against deposition related effects that vary spatially and temporally.  It includes particular emphasis on the indicators and forms of the current standards and the degree to which they are ecologically relevant with regard to deposition related effects. Second, this section relates the current standards and adequacy of protection.
a. Appropriateness
	The ISA has established that the major effects of concern for this review of the oxides of nitrogen and sulfur standards are associated with deposition of N and S caused by atmospheric concentrations of oxides of nitrogen and sulfur.  The current standards are not directed toward depositional effects, and none of the elements of the current NAAQS  -  indicator, form, averaging time, and level  -  are suited for addressing the effects of N and S deposition.
	Four issues arise that call into question the ecological relevance of the structure of the current secondary standards for oxides of nitrogen and sulfur.
(1) The current SO2 secondary standard (0.5 ppm SO2 over a 3-hour average) does not utilize an exposure period that is relevant for ecosystem impacts.  The majority of deposition related impacts are associated with depositional loads that occur over periods of months to years.  This differs significantly from exposures associated with hourly concentrations of SO2 as measured by the current secondary standard.  By addressing short-term concentrations, the current SO2 secondary standard, while protective against direct foliar effects from gaseous oxides of sulfur, does not take into account the findings of effects in the ISA, which notes the relationship between annual deposition of 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 secondary standard. The current secondary standard for oxides of nitrogen is an annual standard (0.053 ppm averaged over 1 year) and as such is more ecologically relevant. Current standards do not utilize appropriate 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 deposition of oxides of nitrogen and sulfur as they do not capture all relevant chemical species of oxidized nitrogen and oxidized sulfur that contribute to deposition.  The ISA provides evidence that deposition related effects are linked with total nitrogen and total sulfur deposition, and thus all forms of oxidized nitrogen and oxidized sulfur that are deposited will contribute to effects on ecosystems.  Thus, by using atmospheric NO2 and SO2 concentrations as indicators, the current standards address only a fraction of total atmospheric oxides of nitrogen and sulfur, and do not take into account the effects from deposition of total atmospheric oxides of nitrogen and sulfur.  This suggests that more comprehensive atmospheric indicators should be considered in designing ecologically relevant standards.
(2) Current standards reflect separate assessments of the two individual pollutants, NO2 and SO2, rather than assessing the joint impacts of deposition to ecosystems.  Recognizing the role that each pollutant plays in jointly affecting ecosystem indicators, functions, and services is vital to developing a meaningful standard.  The clearest example of this interaction is in assessment of the impacts of acidifying deposition on aquatic ecosystems.  Acidification in an aquatic ecosystem depends on the total acidifying potential of the deposition of both N and S from both atmospheric deposition of oxides of nitrogen and sulfur 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 oxides of nitrogen and oxides of sulfur should be a critical element of the conceptual framework for ecologically relevant standards.  There are also important interactions between oxides of nitrogen and sulfur and reduced forms of nitrogen, which also contribute to acidification and nutrient enrichment.  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 oxides of nitrogen and sulfur.  Consideration will also have to be given to total loadings as ecosystems respond to all sources of N and S.
(3) Current standards do not take into account variability in ecosystem sensitivity.  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 or nutrient enrichment as well as atmospheric and landscape characteristics that govern rates of deposition, may lead to standards that do not provide requisite levels of protection across ecosystems.  The current structures of the standards do not address the complexities in the responses of ecosystems to deposition of oxides of nitrogen and sulfur.  Ecosystems contain complex groupings of organisms that respond in various ways to the alterations of soil and water that result from deposition of nitrogen and sulfur compounds.  Different ecosystems therefore respond in different ways depending on a multitude of factors that control how deposition is integrated into the system.  For example, the same levels of deposition falling on limestone dominated soils have a very different effect 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.  Based on the discussed summarized above, the Policy Assessment concludes that the current secondary standards for oxides of nitrogen and oxide sof sulfur are not ecologically relevant in terms of averaging time, form, level or indicator.
a. Adequacy of Protection
	As described in the Policy Assessment, 2005 ambient conditions indicate that the current SO2 and NO2 secondary standards are not exceeded (US EPA, 2011, Figures 6-1 and 6-2) in locations where ecological effects have been observed. In many locations, SO2 and NO2 concentrations are substantially below the levels of the secondary standards.  This pattern suggests that levels of deposition and any effects on ecosystems due to deposition of oxides of nitrogen and sulfur under recent conditions are occurring even though areas meet or are below current standards. In addition, these levels based on conclusions in the REA will not decline in the future to levels below which it is reasonable to anticipate effects.
	 In determining the adequacy of the current secondary standards for oxides of nitrogen and sulfur the Policy Assessment considered the extent to which ambient deposition contributes to loadings in ecosystems.  Since the last review of the secondary standard for oxides of nitrogen, a great deal of information on the contribution of atmospheric deposition associated with ambient oxides of nitrogen has become available.  The REA presents a thorough assessment of the contribution of oxidized nitrogen to nitrogen deposition throughout the U.S., and the relative contributions of ambient oxidized and reduced forms of nitrogen.  The REA concludes that based on that analysis, ambient oxides of nitrogen are a significant component of atmospheric nitrogen deposition, even in areas with relatively high rates of deposition of reduced nitrogen.  In addition, atmospheric deposition of oxidized nitrogen contributes significantly to total nitrogen loadings in nitrogen sensitive ecosystems.
	The ISA summarizes the available studies of relative nitrogen contribution and finds that in much of the U.S., oxides of nitrogen contribute from 50 to 75 percent of total atmospheric deposition relative to total reactive nitrogen that includes oxidized and reduced nitrogen species (US EPA, 2008, section 2.8.4).  Although the proportion of total nitrogen loadings associated with atmospheric deposition of nitrogen varies across locations (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 (US EPA, 2008).
	Atmospheric concentrations of oxides of sulfur 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 S/ha-yr. S was primarily deposited in the form of wet SO4 [2−] followed in decreasing order by a smaller proportion of dry SO2 and a much smaller proportion of deposition as dry SO4[2−].
	As discussed throughout the REA (US EPA, 2009 and section IIB above), there are several key areas of risk that are associated with ambient concentrations of oxides of nitrogen and sulfur.  As noted earlier, in previous reviews of the secondary standards for oxides of nitrogen and sulfur, the standards were designed to protect against direct exposure of plants to ambient concentrations of the pollutants.  A significant shift in understanding of the effects of oxides of nitrogen and sulfur has occurred since the last reviews, reflecting the large amount of research that has been conducted on the effects of deposition of nitrogen and sulfur to ecosystems. The most significant risks of adverse effects to public welfare are those related to deposition of oxides of nitrogen and sulfur to both terrestrial and aquatic ecosystems. These risks fall into two categories, acidification and nutrient enrichment, which were emphasized in the REA are most relevant to evaluating the adequacy of the existing standards in protecting public welfare from adverse ecological effects.
i. Aquatic acidification
	The focus of the REA case studies was on determining whether deposition of sulfur and oxidized nitrogen in locations where ambient oxides of nitrogen and sulfur were at or below the current standards was resulting in acidification and related effects, including episodic acidification and mercury methylation.  Based on the case studies conducted for lakes in the Adirondacks and streams in Shenandoah National Park (case studies are discussed more fully in section IIB and US EPA, 2009), there is significant risk to acid sensitive aquatic ecosystems at atmospheric concentrations of oxides of nitrogen and sulfur at or below the current standards.  The REA also supports strongly a relationship between atmospheric deposition of oxides of nitrogen and sulfur and loss of ANC in sensitive ecosystems and that ANC is an excellent indicator of aquatic acidification.  The REA also concludes that at levels of deposition associated with oxides of nitrogen and sulfur concentrations at or below the current standards, ANC levels are expected to be below benchmark values that are associated with significant losses in fish species richness.
	Significant portions of the U.S. are acid sensitive, and current deposition levels exceed those that would allow recovery of the most acid sensitive lakes in the Adirondacks (US EPA, 2008, Executive Summary).  In addition, because of past loadings, areas of the Shenandoah are sensitive to current deposition levels (US EPA, 2008, Executive Summary).  Parts of the West are naturally less sensitive to acidification and subjected to lower deposition (particularly SOx) levels relative to the eastern United States, and as such, less focus in the ISA is placed on the adequacy of the existing standards in these areas, with the exception of the mountainous areas of the West, which experience episodic acidification due to deposition.
	In describing the effects of acidification in the two case study areas the REA uses the approach of describing benchmarks in terms of ANC values. Many locations in sensitive areas of the U.S. have ANC levels below benchmark levels for ANC classified as severe, elevated, or moderate concern (US EPA, 2011, Figure 2-1). The average current ANC levels across 44 lakes in the Adirondack case study area is 62.1 μeq/L (moderate concern). However, 44 percent of lakes had deposition levels exceeding the critical load for an ANC of 50 μeq/L, and 28 percent of lakes had deposition levels exceeding the critical load for an ANC of 20 μeq/L (US EPA, 2009, section <IP_ADDRESS>). 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 (US EPA, 2009, Figure 4.2-1).  When extrapolated to the full population of lakes in the Adirondacks area using weights based on the EMAP probability survey (US EPA, 2009, section <IP_ADDRESS>), 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 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 (US EPA, 2009, section <IP_ADDRESS>).
	In addition to these chronic acidification effects, the ISA notes that "consideration of episodic acidification greatly increases the extent and degree of estimated effects for acidifying deposition on surface waters." (US EPA, 2008, section <IP_ADDRESS>)  Some studies show that the number of lakes that could be classified as acid-impacted based on episodic acidification is 2 to 3 times the number of lakes classified as acid-impacted based on chronic ANC.  These episodic acidification events can have long term effects on fish populations (US EPA, 2008, section <IP_ADDRESS>).  Under recent conditions, episodic acidification has been observed in locations in the eastern U.S. and in the mountainous western U.S. (US EPA, 2008, section <IP_ADDRESS>).
	The ISA, REA and Policy Assessment all conclude that the current standards are not adequate to protect against the adverse impacts of aquatic acidification on sensitive ecosystems.  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 (US EPA, 2008, Executive Summary). In addition, because of past loadings, areas of the Shenandoah are sensitive to current deposition levels (US EPA, 2008, Executive Summary). A recent survey, as reported in the ISA, found sensitive streams in many locations in the U.S., including the Appalachian Mountains, the Coastal Plain, and the Mountainous West (US EPA, 2008, section <IP_ADDRESS>). In these sensitive areas, between 1 and 6 percent of stream kilometers are chronically acidified. The REA further concludes that both the Adirondack and Shenandoah case study areas are currently receiving deposition from ambient oxides of nitrogen and sulfur in excess of their ability to neutralize such inputs.  In addition based on the current emission scenarios, forecast modeling out to the year 2020 as well as 2050 indicates a large number of streams in these areas will still be adversely impacted (section IIB).  Based on these considerations, the Policy Assessment concludes that the current secondary NAAQS for oxides of nitrogen and sulfur do not provide adequate protection of sensitive ecosystems with regard to aquatic acidification.
ii. Terrestrial acidification
	Based on the terrestrial acidification case studies (Kane Experimental Forest in Pennsylvania and Hubbard Brook Experimental Forest (case study locations described in section IIB) on sugar maple and red spruce habitat, the REA 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.    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 oxides of nitrogen and sulfur and BC:Al, and that BC:Al is a good indicator of terrestrial acidification.  At levels of deposition associated with oxides of nitrogen and sulfur concentrations at or below the current standards, BC:Al levels are expected to be below benchmark values that are associated with significant effects on  tree health and growth. Such degradation of terrestrial ecosystems could affect ecosystem services such as habitat provisioning, endangered species, goods production (timber, syrup, etc.) 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 (US EPA, 2009, section <IP_ADDRESS>).
	Over the full range of sugar maple, 12 percent of evaluated forest plots exceeded the critical loads for a BC:Al ratio of 1.2, and 3 percent exceeded the critical load for a BC:Al ratio of 0.6.  However, there was large variability across states.  In New Jersey, 67 percent of plots exceeded the critical load for a 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, REA and Policy Assessment all conclude that the current standards are not adequate to protect against the adverse impacts of terrestrial acidification on sensitive ecosystems. The ISA reported one study 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 (US EPA, 2008, section <IP_ADDRESS>).  The main drawback, as stated in the REA and Policy Assessment, with the understanding of terrestrial acidification lies in the availability of data by which we can predict critical loads and that the data are based on laboratory responses rather than field measurements. Other stressors that are present in the field but that are not present in the laboratory may confound this relationship. The REA does however, conclude that the case study results suggest that nitrogen and sulfur acidifying deposition in the sugar maple and red spruce forest areas studied were very close to, if not greater than, the critical loads for those areas, and both ecosystems are likely to be sensitive to any future changes in the levels of deposition.
iii. Terrestrial nutrient enrichment
	Nutrient enrichment effects are due to nitrogen loadings from both atmospheric and non-atmospheric sources. Evaluation of nutrient enrichment effects requires an understanding that nutrient inputs are essential to ecosystem health and that specific long term levels of nutrients in a system affect the types of species that occur over long periods of time.  Short term additions of nutrients can affect species competition, and even small additions of nitrogen in areas that are traditionally nutrient poor can have significant impacts on productivity as well as species composition.    Most ecosystems in the United States are nitrogen-limited, so regional decreases in emissions and deposition of airborne nitrogen compounds could lead to some decrease in growth of the vegetation that surrounds the targeted aquatic system but as discussed below evidence for this is mixed. Whether these changes in plant growth are seen as beneficial or adverse will depend on the nature of the ecosystem being assessed.
	Information on the effects of changes in nitrogen deposition on forestlands and other terrestrial ecosystems is very limited. The multiplicity of factors affecting forests, including other potential stressors such as ozone, and limiting factors such as moisture and other nutrients, confound assessments of marginal changes in any one stressor or nutrient in forest ecosystems.  The ISA notes that only a fraction of the deposited nitrogen is taken up by the forests, most of the nitrogen is retained in the soils (US EPA, 2008, section <IP_ADDRESS>). 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 (US EPA, 2008, 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 (US EPA, 2008, section 3.3.9). Because leaching of nitrate can promote cation losses, which in some cases create nutrient imbalances, slower growth and lessened disease and freezing tolerances for forest trees, the net effect of increased N on forests in the U.S. is uncertain (US EPA, 2008, section 3.3.9).
	The scientific literature has many examples of the deleterious effects caused by excessive nitrogen loadings to terrestrial systems.  Several studies have set benchmark values for levels of N deposition at which scientifically adverse effects are known to occur.  Large areas of the country appear to be experiencing deposition above these benchmarks.  The ISA presents that studies have found that at 3.1 kg N/ha-yr, the community of lichens begins to change from acidophytic to tolerant species; at 5.2 kg N/ha-yr, the typical dominance by acidophytic species no longer occurs; and at 10.2 kg N/ha-yr, acidophytic lichens are totally lost from the community. Additional studies in the Colorado Front Range of the Rocky Mountain National Park support these findings. These three values (3.1, 5.2, and 10.2 kg/ha-yr) are one set of ecologically meaningful benchmarks for the mixed conifer forest (MCF) of the pacific coast regions. Nearly all of the known sensitive communities receive total nitrogen deposition levels above the 3.1 N kg/ha-yr ecological benchmark according to 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.  Studies have investigated the amount of nitrogen utilized by healthy and degraded CSS systems. In healthy stands, the authors estimated that 3.3 kg N/ha-yr was used for CSS plant growth. It is assumed that 3.3 kg N/ha-yr is near the point where nitrogen is no longer limiting in the CSS community. Therefore, this amount can be considered an ecological benchmark for the CSS community. The majority of the known CSS range is currently receiving deposition in excess of this benchmark.  Thus, the REA concludes that recent conditions where oxides of nitrogen ambient concentrations are at or below the current oxides of nitrogen secondary standards are not adequate to protect against anticipated adverse impacts from N nutrient enrichment in sensitive ecosystems.
iv. Aquatic nutrient enrichment
	The REA aquatic nutrient enrichment case studies focused on coastal estuaries and revealed that while current ambient loadings of atmospheric oxides of nitrogen are contributing to the overall depositional loading of coastal estuaries, other non-atmospheric sources are contributing in far greater amounts in total, although atmospheric contributions are as large as some other individual source types.  The ability of current data and models to characterize the incremental adverse impacts of nitrogen deposition is limited, both by the available ecological indicators, and by the inability to attribute specific effects to atmospheric sources of nitrogen.  The REA case studies used 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.
	Based on the above considerations, the REA concludes that the ASSETS EI is not an appropriate ecological indicator for estuarine aquatic eutrophication and that additional analysis is required to develop an appropriate indicator for determining the appropriate levels of protection from N nutrient enrichment effects in estuaries related to deposition of oxides of nitrogen.  As a result, EPA is unable to make a determination as to the adequacy of the existing secondary oxides of nitrogen standard in protecting public welfare from 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/ha-yr which are well within current ambient levels are known to cause changes in species composition in diatom communities indicating impaired water quality (US EPA, 2008, section <IP_ADDRESS>).  It then seems apparent then that the existing secondary standard for oxides of nitrogen does not protect such ecosystems and their resulting services from impairment.
v. Other effects
	An important consideration in looking at the effects of deposition of oxides of sulfur in aquatic ecosystems is the potential for production of methylmercury, a highly neurotoxic contaminant.  The production of meaningful amounts of methylmercury (MeHg) requires the presence of SO4[2-] and mercury, and where mercury is present, increased availability of SO4[2-] 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 (US EPA, 2008, section 3.4). In watersheds where changes in sulfate deposition did not produce an effect, one or several of those interacting factors were not in the range required for meaningful methylation to occur (US EPA, 2008, section 3.4). Watersheds with conditions known to be conducive to mercury methylation can be found in the northeastern United States and southeastern Canada (US EPA, 2009, section 6).
	With respect to sulfur deposition and mercury methylation, the final ISA determined: The evidence is sufficient to infer a causal relationship between sulfur deposition and increased mercury methylation in wetlands and aquatic environments. However, EPA did not conduct a quantitative assessment of the risks associated with increased mercury methylation under current conditions. As such, EPA is unable to make a determination as to the adequacy of the existing SO2 secondary standards in protecting against welfare effects associated with increased mercury methylation.
vi. Summary of Adequacy Considerations
	In summary, the Policy Assessment concludes that currently available scientific evidence and assessments clearly call into question the adequacy of the current standards with regard to deposition-related effects on sensitive aquatic and terrestrial ecosystems, including acidification and nutrient enrichment.  Further, the Policy Assessment recognizes that the elements of the current standards -- indicator, averaging time, level and form  -  are not ecologically relevant, and are thus not appropriate for standards designed to provide such protection.  Thus, the Policy Assessment concludes that consideration should be given to establishing a new ecologically relevant multi-pollutant, multimedia standard to provide appropriate protection from deposition-related ecological effects of oxides of nitrogen and sulfur on sensitive ecosystems, with a focus on protecting against adverse effects associated with acidifying deposition in sensitive aquatic ecosystems.
1. CASAC Views
	In a letter to the Administrator (Russell and Samet 2011a), the CASAC Oxides of Nitrogen and Oxides of Sulfur Panel, with full endorsement of the chartered CASAC, unanimously concluded that:
      EPA staff has demonstrated through the Integrated Science Assessment (ISA), Risk and Exposure Characterization (REA) and the draft PA that ambient NOx and SOx can have, and are having, adverse environmental impacts. The Panel views that the current NOx and SOx secondary standards should be retained to protect against direct adverse impacts to vegetation from exposure to gas phase exposures of these two families of air pollutants.  Further, the ISA, REA and draft PA demonstrate that adverse impacts to aquatic ecosystems are also occurring due to deposition of NOx and SOx. Those impacts include acidification and undesirable levels of nutrient enrichment in some aquatic ecosystems. The levels of the current NOx and SOx secondary NAAQS are not sufficient, nor the forms of those standards appropriate, to protect against adverse depositional effects; thus a revised NAAQS is warranted.

	 In addition, with regard to the joint consideration of both oxides of nitrogen and oxides of sulfur as well as the consideration of deposition related effects, CASAC concluded that the Policy Assessment had developed a credible methodology for considering such effects.  The Panel stated that "the Policy Assessment develops a framework for a multi-pollutant, multimedia standard that is ecologically relevant and reflects the combined impacts of these two pollutants as they deposit to sensitive aquatic ecosystems."
2. Administrator's Proposed Conclusions Concerning Adequacy of Current Standard
	Based on the above considerations and taking into account CASAC advice, the Administrator recognizes that the purpose of the secondary standard is to protect against "adverse" effects resulting from exposure to oxides of nitrogen and sulfur, discussed above in section IIA. The Administrator also recognizes the need for conclusions both to the adequacy of the current standards for both direct and deposition related effects as well as conclusions as to the appropriateness and ecological relevance of the current standards.
	In considering what constitutes an ecological effect that is also adverse to the public welfare, the Administrator took into account the ISA conclusions regarding the nature and strength of the effects evidence, the risk and exposure assessment results, the degree to which the associated uncertainties should be considered in interpreting the results, the conclusions presented in the Policy Assessment, and the views of CASAC and members of the public.  On these bases, the Administrator concludes that the current secondary standards are adequate to protect against direct phytotoxic effects on vegetation.  Thus the Administrator proposes to retain the current secondary standards for oxides of nitrogen at 0.053 ppm as an annual average and the current secondary standards for oxides of sulfur at 0.5 ppm as a 3 hour average.
	With regard to deposition-related effects, the Administrator has first to consider the appropriateness of the structure of the current standards to address ecological effects of concern.    Based on the evidence as well as considering the advice given by CASAC on this matter, the Administrator concludes that the elements of the current standards are not ecologically relevant and thus are not appropriate to provide protection of ecosystems.	On the subject of adequacy of protection with regard to deposition-related effects, the Administrator considered the full nature of ecological effects related to the deposition of ambient oxides of nitrogen and sulfur into sensitive ecosystems across the country.  Her conclusions are based on the evidence presented in the ISA with regard to acidification and nutrient enrichment effects, the findings of the REA with regard to scope and severity of the current and likely future effects of deposition, the synthesis of both the scientific evidence and risk and exposure results in the Policy Assessment as to the adequacy of the current standards, and the advice of both CASAC and the public.   After such consideration, the Administrator concludes that current levels of oxides of nitrogen and sulfur are sufficient to cause acidification of both aquatic and terrestrial ecosystems, nutrient enrichment of terrestrial ecosystems and contribute to nutrient enrichment effects in estuaries that could be considered adverse, and the current secondary standards do not provide adequate protection from such effects.
	Having reached these conclusions, the Administrator determines that it is appropriate to consider alternative standards that are ecologically relevant.  These considerations support the conclusion that the current secondary standards is neither appropriate nor adequate to protect against deposition related effects.  The Administrator's consideration of such alternative standards is discussed below in Section III.
III.	Rationale for Proposed Decision on Alternative Secondary Standards for Aquatic Acidification
	Having reached the conclusion that the current NO2 and SO2 secondary standards are not adequate to provide appropriate protection against deposition-related effects associated with oxides of nitrogen and sulfur, the Administrator then considered what new standard might be appropriate, at this time, to address such effects on public welfare.  The Administrator recognizes that the inherently complex and variable linkages between ambient concentrations of nitrogen and sulfur oxides, the related deposited forms of nitrogen and sulfur, and the ecological responses that are associated with public welfare effects call for consideration of an ecologically relevant design of a standard that reflects these linkages.  The Administrator also recognizes that characterization of such complex and variable linkages will necessarily require consideration of information and analyses that have important limitations and uncertainties.
	Despite its complexity, an ecologically relevant multi-pollutant standard to address deposition-related effects could still appropriately be defined in terms of the same basic elements that are used to define any NAAQS  -  indicator, form, averaging time, and level.  The form would incorporate additional structural elements that reflect relevant multi-pollutant and multimedia attributes.  These structural elements include the use of an ecological indicator, tied to the ecological effect we are focused on, and other elements that account for ecologically relevant factors other than ambient air concentrations.  All of these elements would be needed to enable a linkage from ambient air indicators to the ecological indicator to define an ecologically relevant standard.  As a result, such a standard would necessarily be more complex than the NAAQS that have been set historically to address effects associated with ambient concentrations of a single pollutant.
	More specifically, the Administrator considered an ecologically relevant standard to address effects associated with acidifying deposition related to ambient concentrations of oxides of nitrogen and sulfur in sensitive aquatic ecosystems.  This focus is consistent with the information presented in the ISA, REA, and Policy Assessment which highlighted the sufficiency of the quantity and quality of the available evidence and assessments associated with aquatic acidification relative to the information and assessments available for other deposition-related effects, including terrestrial acidification and aquatic and terrestrial nutrient enrichment.  Based on its review of these documents, CASAC agreed that aquatic acidification should be the focus for developing a new multi-pollutant standard in this review.  In reaching conclusions about an air quality standard designed to address deposition-related aquatic acidification effects, the Administrator also recognizes that such a standard may also provide some degree of protection against other deposition-related effects.
	As discussed in chapter 7 of the Policy Assessment, the development of a new standard to address deposition-related aquatic acidification effects recognizes the need for consideration of a nationally applicable standard for protection against adverse effects of aquatic acidification to public welfare, while recognizing the complex and heterogeneous interactions between ambient air concentrations of nitrogen and sulfur oxides, the related deposition of nitrogen and sulfur, and associated ecological responses.  The development of such a standard also needs to take into account the limitations and uncertainties in the available information and analyses upon which characterization of such interactions are based.  The approach used in the Policy Assessment also recognizes that while such a standard would be national in scope and coverage, the effects to public welfare from aquatic acidification will not occur to the same extent in all locations in the U.S., given the inherent variability of the responses of aquatic systems to the effects of acidifying deposition.
	As discussed above in section II, many locations in the U.S. are naturally protected against acid deposition due to underlying geological conditions.  Likewise, some locations in the U.S., including lands managed for commercial agriculture and forestry, are not likely to be negatively impacted by current levels of nitrogen and sulfur deposition.  As a result, while a new ecologically relevant secondary standard would apply everywhere, it would be structured to account for differences in the sensitivity of ecosystems across the country.  This would allow for appropriate protection of sensitive aquatic ecosystems, which are relatively pristine and wild and generally in rural areas, and the services provided by such sensitive ecosystems, without requiring more protection than is needed elsewhere.
	As discussed below, the standard developed in the Policy Assessment would employ (1) NOy and SOx as the atmospheric ambient air indicators; (2) a form that takes into account variable factors, such as atmospheric and ecosystem conditions that modify the amounts of deposited nitrogen and sulfur; the distinction between oxidized and reduced forms of nitrogen; effects of deposited nitrogen and sulfur on aquatic ecosystems in terms of the ecological indicator ANC; and the representativeness of water bodies within a defined spatial area; (3) a multi-year averaging time, and (4) a standard level defined in terms of a single, national target ANC value that, in the context of the above form, identifies the levels of concentrations of NOy and SOx in the ambient air that would meet the standard.  The form of such a standard has been defined by an index, termed an aquatic acidification index (AAI), which reflects the relationship between ambient concentrations of NOy and SOx and aquatic acidification effects that result from nitrogen and sulfur deposition related to these ambient concentrations.
	In presenting the considerations associated with such an air quality standard to address deposition-related aquatic acidification effects, the following sections focus on each element of the standard, including indicator (section III.A), form (section III.B), averaging time (section III.C), and level (section III.D).   Alternative combinations of levels and forms are discussed in section III.E.  Considerations related to important uncertainties inherent in such an approach are discussed in section III.F.  Advice from CASAC on such a new standard is presented in section III.G.  The Administrator's proposed decisions on such a new standard are presented in section III.H.
A.	Ambient Air Indicators
	In considering alternative ambient air indicators, the Policy Assessment primarily focuses on the important attribute of association.  Association in a broad sense refers to how well an ambient air indicator relates to the ecological effects of interest by virtue of both the framework that links the ambient indicator and effects and the empirical evidence that quantifies the linkages.  The Policy Assessment also considers how measurable or quantifiable an indicator is to enable its use as an effective indicator of relevant ambient air concentrations.
	As discussed above in section II.C, the Policy Assessment concludes that indicators other than NO2 and SO2 should be considered as the appropriate indicators of oxides of nitrogen and sulfur in the ambient air for protection against the acidification effects associated with deposition of the associated nitrogen and sulfur.  This conclusion is based on the recognition that all forms of nitrogen and sulfur in the ambient air contribute to deposition and resulting acidification, and as such, NO2 and SO2 are incomplete indicators.  In principle, the ambient indicators should represent the species that are associated with oxides of nitrogen and sulfur in the ambient air and can contribute acidifying deposition.  This includes both the species of oxides of nitrogen and sulfur that are directly emitted as well as species transformed in the atmosphere from oxides of nitrogen and sulfur that retain the nitrogen and sulfur atoms from directly emitted oxides of nitrogen and sulfur.   All of these compounds are associated with oxides of nitrogen and sulfur in the ambient air and can contribute to acidifying deposition.
	The Policy Assessment focuses in particular on the various compounds with nitrogen or sulfur atoms that are associated with oxides of nitrogen and sulfur, because the acidifying potential is specific to nitrogen and sulfur, and not other atoms (e.g., H, C, O) whether derived from the original source of oxides of nitrogen and sulfur emissions or from atmospheric transformations.  For example, the acidifying potential of each molecule of NO2, NO, HNO3 or PAN is identical, as is the potential for each molecule of SO2 or ion of particulate sulfate, p-SO4.   Each atom of sulfur affords twice the acidifying potential of each atom of nitrogen.
1.	Oxides of Sulfur
	As discussed in the Policy Assessment (US EPA, 2011, section 7.1.1), oxides of sulfur include the gases sulfur monoxide (SO), sulfur dioxide (SO2), sulfur trioxide (SO3), disulfur monoxide (S2O), and particulate-phase sulfur compounds that result from gas-phase sulfur oxides interacting with particles.  However, the sum of SO2 and SO4 does represent virtually the entire ambient air mass of sulfur that contributes to acidification.  In addition to accounting for virtually all the potential for acidification from oxidized sulfur in the ambient air, there are reliable methods to monitor the concentrations of SO2 and particulate SO4.  In addition, much of the data used to develop the technical basis for the standard developed in the Policy Assessment is based on monitoring or modeling of these species.  The Policy Assessment concludes that the sum of SO2 and SO4, referred to as SOx, are appropriate ambient air indicators of oxides of sulfur because they represent virtually all of the acidification potential of ambient air oxides of sulfur and there are reliable methods suitable for measuring SO2 and SO4.
2.	Oxides of Nitrogen
	As discussed in the Policy Assessment (US EPA, 2011,section 7.1.2), total reactive oxidized nitrogen, NOy, as defined in chapter 2 of the Policy Assessment, incorporates basically all of the oxidized nitrogen species that have acidifying potential and as such, NOy should be considered as an appropriate indicator for oxides of nitrogen.  NOy is an aggregate measure of s NO and NO2 and all of the reactive oxidized products of NO and NO2.  That is, NOy is a group of nitrogen compounds in which all of the compounds are either an oxide of nitrogen or compounds in which the nitrogen atoms came from oxides of nitrogen.  NOy is especially relevant as an ambient indicator for acidification in that it both relates to the oxides of nitrogen in the ambient air and also represents the acidification potential of all oxidized nitrogen species in the ambient air, whether an oxide of nitrogen or derived from oxides of nitrogen.
	There are currently available reliable methods of measuring aggregate NOy.  The term "aggregate" measure means that the NOY, as measured, is not based on measuring each individual species of NOy and calculating an NOy value by summing the individual species.   Rather, as described in chapter 2 of the Policy Assessment, current measurement techniques process all of the individual NOy species to produce a single aggregate measure of all of the nitrogen atoms associated with any NOy species.  Consequently, the NOy measurement effectively provides the sum of all individual species, but the identity of the individual species is lost.  As discussed above, the accounting for the individual nitrogen atoms is an accounting of the ambient air acidification potential of oxides of nitrogen and their transformation products and therefore the most relevant ambient indicator for aquatic acidification effects associated with oxides of nitrogen.
	This loss of the information on individual species motivated consideration of alternative or more narrowly defined indicators for oxides of nitrogen in the Policy Assessment.  Consideration of a subset of NOy species was based on the following reasoning.  First, the actual dry deposition of nitrogen is determined on an individual species basis by multiplying the species concentration times a species-specific deposition velocity and then summed to develop an estimate of total dry deposition.  Consequently, the use of individual ambient species has the potential to be more consistent with the underlying science of deposition and, therefore, has the potential to allow for a more rigorous evaluation of dry deposition with specialized field studies.  In addition, there has been a suggestion of focusing only on the most quickly depositing NOy species, such as HNO3, as contributions from other NOy species such as NO2 may be negligible.  These alternative indicators are discussed below.
	The Policy Assessment considers the relative merits of using each individual NOy species as part of a group of indicators.  In so doing, it was first noted that dry deposition of NOy is treated as the sum of the deposition of each individual species in advanced process-based air quality models like CMAQ, as described in chapter 2 of the Policy Assessment.  Conceptually one could extend this process-based approach by using all NOY species individually as separate indicators for oxides of nitrogen and requiring, for example, measurements of each of the species, including the dominant species of HNO3, particulate nitrate (p-NO3), true NO2, NO, and PAN.   The potential attraction of using individual species would be the reliance on actual deposition velocities.  This could have more physical meaning in comparison to a constructed model of aggregate deposition of NOy, which is difficult to evaluate with observations because of the assimilation of many species with disparate deposition behavior.  The Policy Assessment notes that the major drawback of using individual NOy species as the indicators is the lack of reliable measurement techniques, especially for PAN and NO2 in rural locations, which renders the use of virtually any individual NOy species, except for NO and perhaps p-NO3, as functionally inadequate from a measurement perspective.
	The Policy Assessment next considered the relative merits of using a subset of NOy species as the indicators for oxides of nitrogen, as was discussed above for oxides of sulfur.  To the extent that certain species provide relatively minor contributions to total NOy deposition, it may be appropriate to consider excluding them as part of the indicator.  As discussed in chapter 2 of the Policy Assessment, each nitrogen species within the array of NOy species has species-specific dry deposition velocities.   For example, the deposition velocity of HNO3 is much greater than the velocity for nitrogen dioxide and, consequently, for a similar ambient air concentration, HNO3 contributes more deposition of acidifying nitrogen relative to nitrogen dioxide.  In transitioning from source-oriented urban locations to rural environments, the ratio of the concentrations of HNO3 and PAN to NO2 increases.
	Based on the reasoning that a larger fraction of the deposited NOY is accounted for by total nitrate (the sum of HNO3 and p-NO3), a surrogate for the more rapidly depositing fraction of NOy, combined with the availability of reliable total nitrate measurements through the Clean Air Status and Trends Network (CASTNET), the Policy Assessment considered using total nitrate as the indicator for oxides of nitrogen (US EPA, 2011, appendix E).  Nitrate would be expected to correlate well with total reactive oxidized nitrogen deposition relative to NOY (US EPA, 2011,chapter 2), given the inherent noise associated with variable contributions of low deposition velocity species (e.g., NO2) that may have relatively high ambient concentrations.  However, modeling simulations suggest that NOy may be a more robust indicator, relative to HNO3, in terms of relating absolute changes in ambient air concentrations to changes in nitrogen deposition driven by changes in ambient concentrations of oxides of nitrogen (US EPA, 2011,Figure 2-32).
	Based on the above considerations, the Policy Assessment concludes that total reactive oxidized nitrogen, NOy, should be considered as the appropriate ambient indicator for oxides of nitrogen based on its direct relationship to oxides of nitrogen in the ambient air and its direct relationship to deposition associated with aquatic acidification.  Because NOy represents all of the potentially acidifying oxidized nitrogen species in the ambient air, it is appropriately associated with the deposition of potentially acidifying compounds associated with oxides of nitrogen in the ambient air.  In addition, there are reliable methods available to measure NOy.  Measurement of each individual species of NOy, or the measurement of only a subset of species of NOy, is less appropriate because there are not reliable measurements methods available to measure all of the individual species of NOy and a subset of species would fail to account for significant portions of the oxidized reactive nitrogen that relates to acidification.
B.	Form
	Based on the evidence of the aquatic acidification effects caused by the deposition of NOy and SOx, the Policy Assessment (US EPA, 2011,section 7.2) presents the development of a new form that is ecologically relevant for addressing such effects.  The conceptual design for the form of such a standard includes three main components:  an ecological indicator, deposition metrics that relate to the ecological indicator, and a function that relates ambient air indicators to deposition metrics.  Collectively, these three components link the ecological indicator to ambient air indicators, as illustrated above in Fig II-1.
	The simplified flow diagram in Figure II-1 compresses the various atmospheric, biological, and geochemical processes associated with acidifying deposition to aquatic ecosystems into a simplified conceptual picture.  The ecological indicator (left box) is related to atmospheric deposition through biogeochemical ecosystem models (middle box), which associate a target deposition load to a target ecological indicator.   Once a target deposition is established, associated allowable air concentrations are determined (right box) through the relationships between concentration and deposition that are embodied in air quality models such as CMAQ.   The following discussion describes the development and rationale for each of these components, as well as the integration of these components into the full expression of the form of the standard using the concept of a national AAI that represents a target ANC level as a function of ambient air concentrations.  Spatial aggregation issues associated with defining each of the terms of this index are also addressed below.
 	The AAI is designed to be an ecologically relevant form of the standard that determines the levels of NOy and SOx in the ambient air that would achieve a target ANC limit for the U.S.  The intent of the AAI is to weight atmospheric concentrations of oxides of nitrogen and sulfur by their propensity to contribute to acidification through deposition, given the fundamental acidifying potential of each pollutant, and to take into account the ecological factors that govern acid sensitivity in different ecosystems.  The index also accounts for the contribution of reduced nitrogen to acidification.  Thus, the AAI encompasses those attributes of specific relevance to protecting ecosystems from the acidifying potential of ambient air concentrations of NOy and SOx.
1.	Ecological Indicator
 	In considering alternative ecological indicators, the Policy Assessment again primarily focuses on the attribute of association.  In the case of an ecological indicator for aquatic acidification, association refers to the relationship between the indicator and adverse effects as discussed in section II.  Because of the conceptual structure of the form of this standard (Figure III-1), this particular ecological indicator must also link up in a meaningful and quantifiable manner with acidifying atmospheric deposition.  In effect, the ecological indicator for aquatic acidification is the bridge between biological impairment and deposition of NOy and SOx.
	This section presents the rationale in the Policy Assessment for selecting acid neutralizing capacity (ANC) as the appropriate ecological indicator for consideration.  Recognizing that ANC is not itself the causative or toxic agent for adverse aquatic acidification effects, the rationale for using ANC as the relevant ecological indicator is based on the following:
   (1) ANC is directly associated with the causative agents, pH and dissolved aluminum, both through empirical evidence and mechanistic relationships;
   (2) Empirical evidence shows very clear and strong relationships between adverse effects and ANC;
   (3) ANC is a more reliable indicator from a modeling perspective, allowing use of a body of studies and technical analyses related to ANC and acidification to inform the development of the standard; and
   (4) ANC literally embodies the concept of acidification as posed by the basic principles of acid base chemistry and the measurement method used to estimate ANC and, therefore, serves as a direct index to protect against acidification.

	Ecological indicators of acidification in aquatic ecosystems can be chemical or biological components of the ecosystem that are altered by the acidifying effects of nitrogen and sulfur deposition. A desirable ecological indicator for aquatic acidification is one that is measurable or estimable, linked causally to deposition of nitrogen and sulfur, and linked causally, either directly or indirectly to ecological effects known or anticipated to adversely affect public welfare.
	As summarized in chapter 2 of the Policy Assessment, atmospheric deposition of NOy and SOx causes aquatic acidification through the input of strong acid anions (e.g., NO3[-] and SO4[2-]) that ultimately shifts the water chemistry equilibrium toward increased hydrogen ion levels (or decreased pH).  The anions are deposited either directly to the aquatic ecosystem or indirectly via transformation through soil nitrification processes and subsequent drainage from terrestrial ecosystems.  In other words, when these anions are mobilized in the terrestrial soil, they can leach into adjacent water bodies.  Aquatic acidification is indicated by changes in the surface water chemistry of ecosystems. In turn, the alteration of surface water chemistry has been linked to negative effects on the biotic integrity of freshwater ecosystems. There is a suite of chemical indicators that could be used to assess the effects of acidifying deposition on lake or stream acid-base chemistry. These indicators include acid neutralizing capacity (ANC); alkalinity (ALK); base neutralizing capacity, commonly referred to as acidity (ACY); surface water pH; concentrations of trivalent aluminum, Al[+3]; and concentrations of major anions (SO4[2-], NO3[-]), cations (Ca[2+], Mg[+2], K[+]), or sums of cations or anions.
	ANC and ALK are very similar quantities and are used interchangeably in the literature and for some of the analyses presented in this document.   Both ANC and ALK are defined as the amount of strong acid required to reach a specified equivalence point.  For acid-base solutions, an equivalence point can be thought of as the point to which the addition of strong acids (i.e., titration) is no longer neutralized by the solution.   This explains the term acid neutralizing capacity, or ANC, as ANC relates directly to the capacity of a system to neutralize acids.  The differences between ANC and ALK are based on operational definitions and subject to various interpretations.   ANC is preferred over ALK as the body of scientific evidence has focused on ANC and effects relationships.   ALK is more widely associated with more general characterizations of water quality such as the relative hardness of water associated with carbonates.
	Indictors such as the concentrations of specific anions, cations, or their groupings, while relevant to acidification processes, are not robust acidification indicators as it is the relative balance of cations and anions that is more directly associated with acidification.  That balance is captured by ANC and ALK.   Acidity, ACY, is the corollary of ANC from the perspective of how much strong base it takes to reach an equivalence point.  Because ACY is not used in most ecosystem assessments, the body of information relating ACY to effects is too limited to serve as a basis for an appropriate ecological indicator.  Aluminum and other metals are causative toxic agents that directly impair biological functions.  However, aluminum, or metals in general, have high variability in concentrations that can be linked to effects, often at extremely low levels which in some cases approach detectability limits, exhibit rapid transient responses, and are often confounded by the presence of other toxic metals. These concerns limit the use of metals as reliable and measurable ecological indicators.  Hydrogen ion (H[+]) concentrations, using their negative logarithmic values, or pH, are well correlated with adverse effects, as discussed above in section II.A, and determine the solubility of metals such as aluminum.  However, pH is not a preferred acidification indicator due to its highly transient nature and other concerns, as discussed below.
	Having reasoned that ANC is a preferred indicator to ALK, ACY, individual metals or groupings of ions, the Policy Assessment considers the relative merits of ANC compared to pH, which is a well recognized indicator of acidity and a more direct causative agent with regard to adverse effects.  First, the linkage between ANC and pH is considered in recognition of the causative association between pH and effects.
	ANC is not the direct causative toxic agent impacting aquatic species diversity.   The scientific literature generally emphasizes the links between pH and adverse effects as described above in section II.A.   It is important, therefore, to consider the extent to which ANC and pH are well related from a mechanistic perspective as well as through empirical evidence.   ANC and pH are co-dependent on each other based on the requirement that all solutions are electrically neutral, meaning that any solution must satisfy the condition that all negatively charged species must be balanced by all positively charged species.   ANC is defined as the difference between strong anions and cations (US EPA, 2011, equation 7-13).
	While the chemistry can be complex, the co-dependency between ANC and pH is explained by recognizing that positively charged hydrogen, H[+], is incorporated in the charge balance relationships related to the overall solution chemistry which also defines ANC.  The positive, directional co-dependency (i.e., ANC and pH increase together) is further explained in concept as ANC reflects how much strong acid (i.e., how much hydrogen ion) it takes to titrate to an equivalence point.  Strong observed correlations between pH and ANC as described in the Policy Assessment support these mechanistic relationships.
	As discussed above in section II.A, there are well established examples of ANC correlating strongly with a variety of ecological effects which are summarized in the Policy Assessment (US EPA, 2011, Table 3-1).   Because pH and ANC are well correlated and linearly dependent over the pH ranges (4.5-6) where adverse ecological effects are observed, evidence of clear associations exist between ANC and adverse ecological effects as described in the Policy Assessment.  In large measure, this dependence between pH and ANC and the relationship of both pH and ANC to effects, speak directly to the appropriateness of ANC with respect to its use as an ecological indicator.
	Thus, there is a clear association between ANC and ecological effects, although there is a more direct causal relationship between pH and effects.  Nonetheless, ANC is preferred as an ecological indicator based on its superior ability to provide a linkage with deposition in a meaningful and quantifiable manner, a role that is served far more effectively by ANC than by pH.  While both ANC and pH are clearly associated with the effects of concern, ANC is superior in linking these effects to deposition.
	The Policy Assessment notes that the basis for this conclusion is that acidifying atmospheric deposition of nitrogen and sulfur is a direct input of potential acidity (ACY), or, in terms of ANC, such deposition is relevant to the major anions that reduce the capacity of a water body to neutralize acidity.  Consequently, there is a well defined linear relationship between potential acidifying deposition and ANC.  This ANC-deposition relationship facilitates the linkage between ecosystem models that calculate an ecological indicator and the atmospheric deposition of NOy and SOx.  On the other hand, there is no direct linear relationship between deposition and pH.  While acid inputs from deposition lower pH, the relationship can be extremely nonlinear and there is no direct connection from a modeling or mass balance perspective between the amount of deposition entering a system and pH.  The term "mass balance" underlies the basic formulation of any physical modeling construct, atmospheric or aquatic systems, and refers to the accounting of the flow of mass into a system, the transformation to other forms, and the loss due to flow out of a system and other removal processes.  ANC is a conserved property.  This means that ANC in a water body can be accounted for by knowledge of how much ANC initially exists, how much flows in and is deposited, and how much flows out.  In contrast, hydrogen ion concentration in the water, the basis for pH, is not a conserved property as its concentration is affected by several factors such as temperature, atmospheric pressure, mixing conditions of a water body, and the levels of other several chemical species in the system.  The disadvantage of pH lacking conservative properties is that there is a very complex connection between changes in ambient air concentrations of NOy and SOx and pH.
	The discussion of basic water chemistry of natural systems in chapter 2 of the Policy Assessment provides further details on why pH is not a conserved quantity and is subject to rapid transient response behavior that makes it difficult to use as a reliable and functional ecological indicator.  The observed pH-to-ANC relationship (US EPA, 2011, figure 7-2) partially explains the concern with pH responding too abruptly.   In the region where pH ranges roughly from 4.5 to 6 and is of greatest relevance to effects (US EPA, 2011, figure 7-4), there clearly is more sensitivity of pH to changes in ANC in the ANC range from approximately 0 to 50 ueq/L.   A focus on this part of the ANC-to-pH relationship shows that ANC associates well with pH in a fairly linear manner.  However, the pH range from 4.5 to 6 also includes one of the very steepest parts of the slope relating pH as a function of ANC, where ANC ranges down below 0 ueq/L, which is subject to very rapid change in ANC, or deposition inputs.   This part of the relationship coincides with reduced levels of ANC and hence with reduced ability to neutralize acids and moderate pH fluctuations.  This response behavior can be extended to considering how pH would change in response to deposition, or ambient concentrations, of NOy and SOx, which can be viewed as "ANC-like" inputs.
	In summary, because ANC clearly links both to biological effects of aquatic acidification as well as to acidifying inputs of NOy and SOx deposition, the Policy Assessment concludes that ANC is an appropriate ecological indicator for relating adverse aquatic ecosystem effects to acidifying atmospheric deposition of SOx and NOy, and is preferred to other potential indicators.  In reaching this conclusion, the Policy Assessment notes that in its review of the first draft Policy Assessment, CASAC concluded that "information on levels of ANC protective to fish and other aquatic biota has been well developed and presents probably the lowest level of uncertainty in the entire methodology" (Russell and Samet, 2010a).  In its more recent review of the second draft Policy Assessment, CASAC agreed "that acid neutralizing capacity is an appropriate ecological measure for reflecting the effects of aquatic acidification" (Russell and Samet, 2010b; p. 4).
2.	Linking ANC to Deposition
	There is evidence to support a quantified relationship between deposition of nitrogen and sulfur and ANC.  This relationship was analyzed in the REA for two case study areas, the Adirondack and Shenandoah Mountains, based on time-series modeling and observed trends.
In the REA analysis, long-term trends in surface water nitrate, sulfate and ANC were modeled using Model of Acidification of Groundwater in Catchment (MAGIC) for the two case study areas.  These data were used to compare recent surface water conditions in 2006 with preindustrial conditions (i.e. preacidification 1860).  The results showed a marked increase in the number of acid impacted lakes, characterized as a decrease in ANC levels, since the onset of anthropogenic nitrogen and sulfur deposition, as discussed in chapter 2 of the Policy Assessment.
In the REA, more recent trends in ANC, over the period from 1990 to 2006, were assessed using monitoring data collected at the two case study areas.  In both case study areas, nitrate and sulfate deposition decreased over this time period.  In the Adirondack Mountains, this corresponded to a decreased concentration of nitrate and sulfate in the surface waters and an increase in ANC (US EPA, 2009, section <IP_ADDRESS>).  In the Shenandoah Mountains, there was a slight decrease in nitrate and sulfate concentration in surface waters corresponding to modest increase in ANC from 50 ueq/L in 1990 to 67 ueq/L in 2006 (US EPA, 2009, section <IP_ADDRESS>,Appendix 4, and section 3.4).
	In the REA, the quantified relationship between deposition and ANC was investigated using ecosystem acidification models, also referred to as acid balance models or critical loads models (US EPA, 2011, section 2 and US EPA, 2009, section 4 and Appendix 4).  These models quantify the relationship between deposition of nitrogen and sulfur and the resulting ANC in surface waters based on an ecosystem's inherent generation of ANC and ability to neutralize nitrogen deposition through biological and physical processes.  A critical load is defined as the amount of acidifying atmospheric deposition of nitrogen and sulfur beyond which a target ANC is not reached.   Relatively high critical load values imply that an ecosystem can accommodate greater deposition levels than lower critical loads for a specific target ANC level.   Ecosystem models that calculate critical loads form the basis for linking deposition to ANC.
	As discussed in chapter 2 of the Policy Assessment, both dynamic and steady state models calculate ANC as a function of ecosystem attributes and atmospheric nitrogen and sulfur deposition, and can be used to calculate critical loads.   Steady state models are time invariant and reflect the long term consequences associated with an ecosystem reaching equilibrium under a constant level of atmospheric deposition.   Dynamic models are time variant and take into account the time dependencies inherent in ecosystem hydrology, soil and biological processes.   Dynamic models like MAGIC can provide the time series response of ANC to deposition whereas steady state models provide a single ANC relationship to any fixed deposition level.   Dynamic models naturally are more complex than steady state models as they attempt to capture as much of the fundamental biogeochemical processes as practicable, whereas steady state models depend on far greater parameterization and generalization of processes that is afforded, somewhat, by not having to accounting for temporal variability.
	 The Policy Assessment notes that steady state models are capable of addressing the question of what does it take to reach and sustain a specific level of ANC.  Dynamic models are also capable of addressing that question, but can also address the question of how long it takes to achieve that result.  Dynamic models afford the ability for more comprehensive treatment of a variety of processes throughout the surface, soil and bedrock layers within an ecosystem.  For example, steady state models treat sulfate as a mobile anion throughout the system, meaning that the sulfate that is deposited to a watershed enters the water column and is not influenced by soil adsorption or cation exchange.  Dynamic models can incorporate these time variant processes.  The use of a steady state model treating sulfate as totally mobile does not necessarily conflict with the possibility of sulfate acting as a less than mobile ion at certain times.   The steady state assumption is premised on the long term behavior of sulfate which can undergo periods of net adsorption followed by periods of net desorption which can balance out over time.  The Policy Assessment recognizes that as the richness of the available data increases, in terms of parameters and spatial resolution, the incorporation of dynamic modeling approaches in the standard setting process should become more feasible.  In determining an appropriate modeling approach for the development of a NAAQS in this review, the Policy Assessment considers both the relevance of the question addressed as well as the ability to perform modeling that provides relevant information for geographic areas across the country.
	Dynamic models require a large amount of catchment level-specific data relative to steady state models.  Because of the time invariant nature of steady state models, the data requirements that integrate across a broad spectrum of ecosystem processes is achievable and available now at the national level.  Water quality data to support steady state models currently exist for developing a national data base for modeling nearly 10,000 catchments in the contiguous U.S.  In contrast, the data needs to support dynamic models for national-scale analyses simply are not available at this time.  Further, the information provided by steady state modeling would be sufficient to develop and analyze alternative NAAQS and the kind of protection they would afford.  While it would be of interest to also obtain information about how much time it would take for a target ANC level to be achieved, the absence of such information does not preclude developing and evaluating alternative NAAQS using the AAI structure.  Based on the above considerations, the Policy Assessment concludes that at this time steady state critical load modeling is an appropriate tool for linking long-term ANC levels to atmospheric deposition of nitrogen and sulfur for development of an AAI that has national applicability.
	 A steady state model is used to define the critical load, which is the amount of atmospheric deposition of nitrogen (N) and sulfur (S) beyond which a target ANC is not achieved and sustained.  It is expressed as:

CLANClim(N + S) = ([BC]0[*] - [ANClim])Q + Neco					(III-1)

Where:
 CLANClim(N + S) is the critical load of deposition, with units of equivalent charge/(area-time);
 [BC]0[*] is the natural contribution of base cations from weathering, soil processes and preindustrial deposition, with units of equivalent charge/volume;
[ANClim] is the target ANC value, with units of equivalent charge/volume;
Q is the catchment level runoff rate governed by water mass balance and dominated by precipitation, with units of distance/time; and
Neco is the amount of nitrogen deposition that is effectively neutralized by a variety of biological (e.g., nutrient uptake) and physical processes, with units of equivalent charge/ (area-time).

	Equation III- 1 is a modified expression that adopts the basic formulation of the steady state models that are described in chapter 2 of the Policy Assessment.  More detailed discussion of the rationale, assumptions and derivation of equation III- 1, as well as all of the equations in this section, are included in Appendix B of the Policy Assessment.  The equation simply reflects the amount of deposition of nitrogen and sulfur from the atmosphere, CLANClim(N + S), that is  associated with a sustainable long-term ANC target, [ANClim], given the capacity of the natural system to generate ANC, [BC]0[*], and the capacity of the natural system to neutralize nitrogen deposition, Neco.  This expression of critical load is valid when nitrogen deposition is greater than Neco.  The runoff rate, Q, allows for balancing mass in the two environmental mediums  -  atmosphere and catchment.  This critical load expression can be focused on a single water system or more broadly.  To extend applicability of the critical load expression (equation III-1) from the catchment level to broader spatial areas, the terms Qr and CLr, are used, which are the runoff rate and critical load, respectively, of the region over which all the atmospheric terms in the equation are defined.
	In considering the contributions of SOx or NOy species to acidification, it is useful to think of every depositing nitrogen atom as supplying one equivalent charge unit and every sulfur atom as depositing two charge units.  The Policy Assessment uses equivalent charge per volume as a normalizing tool in place of the more familiar metrics such as mass or moles per volume.  This allows for a clearer explanation of many of the relationships between atmospheric and ecosystem processes that incorporate mass and volume unit conventions somewhat specific to the environmental media of concern (e.g., m[3] for air and liter for liquid water).  Equivalent charge reflects the chemistry equilibrium fundamentals that assume electroneutrality, or balancing charge where the sum of cations always equals the sum of anions.
	As presented above, the terms S and N in the CLANClim (N + S) term broadly represent all species of sulfur or nitrogen that can contribute to acidifying deposition.  This follows conventions used in the scientific literature that addresses critical loads, and it reflects all possible acidifying contributions from any sulfur or nitrogen species.  For all practical purposes, S reflects SOx as described above, the sum of sulfur dioxide gas and particulate sulfate.   However, N in equation III-1 includes both oxidized forms, consistent with the ambient indicator, NOy, in addition to the reduced nitrogen species, ammonia and ammonium ion, referred to as NHx.   NHx is included in the critical load formulation because it contributes to potentially acidifying nitrogen deposition.  Consequently, from a mass balance or modeling perspective, the form of the standard needs to account for NHx, as described below.
3.	Linking Deposition to Ambient Air Indicators
	The last major component of the form illustrated in Figure III-1 addresses the linkage between deposition of nitrogen and sulfur and concentrations of the ambient air indicators, NOy and SOx.   To link ambient air concentrations with deposition, the Policy Assessment defines a transference ratio, T, as the ratio of total wet and dry deposition to ambient concentration, consistent with the area and time period over which the standard is defined.  To express deposition of NOy and SOx in terms of NOy and SOx ambient concentrations, two transference ratios were defined, where TSOx equals the ratio of the combined dry and wet deposition of SOx to the ambient air concentration of SOx, and TNOy equals the ratio of the combined dry and wet deposition of NOy to the ambient air concentration of NOy.
	As described in Chapter 7 of the Policy Assessment, reduced forms of nitrogen (NHX) are included in total nitrogen in the critical load equation, III-1.  Reduced forms of nitrogen are treated separately, as are NOy and SOx, and the transference ratios are applied.  This results in the following critical load expression that is defined explicitly in terms of the indicators NOy and SOx:

CLANClim(N + S) = ([BC]0[*] - [ANClim])Q + Neco = [NOy]TNOy + [SOx]TSOx + NHx	(III-2)

This is the same equation as III-1, with the deposition associated with the critical load translated to deposition from ambient air concentrations via transference ratios.  In addition, deposition of reduced nitrogen, oxidize nitrogen and oxidized sulfur are treated separately.
	Transference ratios are a modeled construct, and therefore cannot be compared directly to measurable quantities.  There is an analogy to deposition velocity, as a transference ratio is basically an aggregated weighted average of the deposition velocities of all contributing species across dry and wet deposition, and transference ratio units are expressed as distance/time.  However, wet deposition commonly is not interpreted as the product of a concentration times a velocity.  Direct wet deposition observations are available which integrate all of the processes, regardless of how well they may be understood, related to wet deposition into a measurable quantity.  There are reasonable analogies between the processes governing dry and wet deposition, from a fundamental mass transfer perspective.  In both cases there is a transfer of mass between the dry ambient phase and another medium, either a surface or vegetation in the case of dry deposition, or a rain droplet or cloud in the case of wet precipitation. The specific thermodynamic properties and chemical/biological reactions that govern the transfer of dry mass to plants or aqueous droplets differ, but either process can be based on conceptualizing the product of a concentration, or concentration difference, times a mass transfer coefficient which is analogous to the basic dry deposition model:  dry deposition = concentration x velocity (US EPA, 2011, Appendix F).
	Transference ratios require estimates of wet deposition of NOy and SOx, dry deposition of NOy and SOx, and ambient air concentrations of NOy and SOx.  Possible sources of information include model estimates or a combination of model estimates and observations, recognizing that dry deposition is a modeled quantity that can use observed or modeled estimates of concentration.  The limited amount of NOy measurements in acid-sensitive areas as well as the combination of representative NOy, SO2 and SO4 observations generally preclude the use of observations for development of a standard that is applicable nationally.
	The Policy Assessment considers a blending of observations and models to take advantage of their relative strengths; e.g., combining the NADP wet deposition observations, modeled dry deposition, and a mix of modeled and observed concentrations, using the model for those species not measured or measured with very sparse spatial coverage.  A potential disadvantage of mixing and matching observations and model estimates is to lose consistency afforded by using just modeling alone.  A modeling platform like CMAQ is based on adhering to consistent treatment of mass conservation, by linking emission inputs with air concentrations and concentrations to deposition.  Inconsistencies from combining processes from different analytical platforms increase the chance that mass (of nitrogen or sulfur) would unintentionally be increased or decreased as the internal checking that assures mass conservation is lost.  	Transference ratios incorporate a broad suite of atmospheric processes and consequently an analytical approach that instills consistency in the linkage of these processes is preferable to an approach lacking such inherent consistency.  This contention does not mean that observations alone, if available, could not be used, but suggests that the inconsistencies in combining models and observations for the purposes of developing transference ratios has the potential for creating unintended artifacts.
	While there is a reasonable conceptual basis for the concept of an aggregated deposition velocity referred to in the Policy Assessment as a transference ratio, there is very limited ability to compare observed and calculated ratios.  This is because the deposition velocity is dependent on individual species, and the mass transfer processes of wet and dry removal, while conceptually similar, are different.  Consequently, there does not exist a meaningful approach to measure such an aggregated or lumped parameter.  Therefore, at this time, the evaluation of transference ratios is based on sensitivity studies, analysis of variability, and comparisons with other models, as described in Appendix F of the Policy Assessment.
	As discussed in Appendix F, the interannual variability, as well as the sensitivity to emission changes of roughly 50%, results in changes of transference ratios of approximately 5 - 10%.  Part of the reason for this inherent stability is due to the co-dependence of concentration and deposition.  For example, as concentrations are reduced as a result of emissions reductions, deposition in turn is reduced since deposition is a direct linear function of concentration leading to negligible impact on the deposition-to-concentration ratio.  Likewise, an overestimate of concentration likely does not induce a bias in the transference ratio.  While it is important to continue to improve the model's ability to match ambient concentrations in time and space, the bias of a modeled estimate of concentration relative to observations does not necessarily result in a bias in a calculated transference ratio.  In effect, this consideration of bias cancellation reduces the sensitivity of transference ratios to model uncertainties and affords increased confidence in the stability of these ratios.  Based on the series of sensitivity and variability analyses, the Policy Assessment concludes that the transference ratios are relatively stable and provide a sound metric for linking deposition and concentration in the form of the standard.
	As discussed in the policy Assessment, transference ratios are dependent on the platform upon which they are constructed.  Comparisons of transference ratios constructed from different modeling platforms do exhibit significant differences.  While this divergence of results may be explained by a variety of differences in process treatments, input fields and incommensurabilities in species definitions and spatial configurations, it does suggest two very important conclusions.  First, the idea of using multiple platforms for different parts of the country may be problematic as there does not exist a reliable approach to judge acceptance which is almost always based on comparisons to observations.  Second, since transference ratios are based on concentrations and deposition, as the uncertainties in each of those components are reduced, the relative uncertainty in the ratios also is reduced.  This means that basic improvements in the model's ability to reproduce observed wet deposition and ambient concentration fields enhance the relative confidence in the constructed transference ratios.  Similarly, as in-situ dry deposition flux measurements become available that enable a more rigorous evaluation and diagnosis of modeled dry deposition processes, the expected improved treatment of dry deposition also would increase confidence in transference ratios.  Finally, deposition is directly related to ambient air concentrations.  Models like CMAQ rely on the concentration-to-deposition linkage to calculate deposition, which is the foundation for broadly based and robust assessments addressing atmospheric deposition.  In principle, the use of a modeled constructed transference ratio is based on the same premise by which we use models to estimate deposition in the first place.
	The shortage of widely available ambient air observations and the fact that estimates of dry deposition requires modeling, collectively suggests that a unified modeling platform is the best approach for constructing transference ratios.  The Policy Assessment (US EPA, 2011,section 2) considers CMAQ and other models, such as CAMx and the Canada's AURAMS - A Unified Regional Air-quality Modeling System (Smythe et al., 2008), and concludes that CMAQ is the preferred modeling platform for constructing transference ratios for purposes of developing a new secondary standard for consideration.  This conclusion reflects the view that for the purposes of defining transference ratios, a modeling platform should (1) be a multiple pollutant model recognizing the myriad of connections across pollutant categories that directly and indirectly impact nitrogen and sulfur characterization, (2) include the most comprehensive scientific treatments of atmospheric processes that relate directly and indirectly to characterizing concentrations and deposition, (3) have an infrastructure capability that accommodates the inclusion of improved scientific treatments of relevant processes and important input fields, and (4) undergo frequent reviews regarding the adequacy of the underlying science as well as the appropriateness in applications.  The CMAQ platform exhibits all these characteristics.  It has been (and continues to be) extensively evaluated for several pollutant categories, is supported by a central infrastructure of EPA scientists, with considerable interfacing with the scientific research communities in academia and industry,  whose mission is to improve and evaluate the CMAQ platform.  More directly, CMAQ, and its predecessor versions, has a long track record going back to the NAPAP in the 1980's of specific improvements in deposition processes, which are described in Appendix F of the Policy Assessment.
4.	Aquatic Acidification Index
	Having established the various expressions that link atmospheric deposition of nitrogen and sulfur to ANC and the transference ratios that translate atmospheric concentrations to deposition of nitrogen and sulfur, the Policy Assessment derives the following expression of these linkages, which separates reduced forms of nitrogen, NHx, from oxidized forms:

ANCcalc = {ANClim + CLr/Qr }  -  NHx/Qr - TNOy [NOy]/Qr - TSOx[SOx]/Qr 		(III-3)

	Equation III- 3 is the basic expression of the form of a standard that translates the conceptual framework into an explicit expression that defines ANC as a function of the ambient air indicators, NOy and SOx.  This equation calculates an expected ANC value based on ambient concentrations of NOy and SOx.  The calculated ANC will differ from the target ANC (ANClim) depending on how much the nitrogen and sulfur deposition associated with NOy, SOx, and NHx differs from the critical load associated with just achieving the target ANC.
Based on equation III- 3, the Policy Assessment defines an aquatic acidification index (AAI) that is more simply stated using terms that highlight the ambient air indicators:

AAI  =  F1  -  F2  -  F3[NOy]  -  F4[SOx]						(III- 4)

where the AAI represents the long term (or steady state) ANC level associated with ambient air concentrations of NOy and SOx.   The factors F1 through F4 convey three attributes:  a relative measure of the ecosystem's ability to neutralize acids (F1), the acidifying potential of reduced nitrogen deposition (F2), and the deposition-to-concentration translators for NOy (F3) and SOx (F4).
Specifically:
F1 = ANClim + CLr/Qr ;
F2 =  NHx/Qr  = NHx deposition divided by Qr;
F3 =  TNOy/ Qr ; TNOy is the transference ratio that converts ambient air concentrations of NOy to deposition of NOy; and
F4 =  TSOx/ Qr ; TSOx is the transference ratio that converts ambient air concentrations of SOx to deposition of SOx.
All of these factors include representative Qr to maintain unit (and mass) consistency between the AAI and the terms on the right side of equation III-4.