# Risk and Exposure Assessment for Review of the Secondary National Ambient Air Quality Standards for Oxides of Nitrogen and Oxides of Sulfur

First Draft, Chapters 7-8[This page intentionally left blank.]EPA-452/P-08-005b August 2008

RISK AND EXPOSURE ASSESSMENT FOR REVI EW OF THE SECONDARY NATIONAL AMBIENT AIR QUALITY STANDARDS FOR OXIDES OF NITROGEN AND OXIDES OF SULFUR
FIRST DRAFT, CHAPTERS 7-8

U.S. Environmental Protection Agency Office of Air Quality Planning and Standards Research Triangle Park, NC

## DISCLAIMER

This draft document has been prepared by staff from the Health and Environmental Impacts and Air Quality An alysis Divisions of the Office of Air Quality Planning and Standards, and th e Clean Air Markets Division, Office of Air Programs, U.S. Environmental Protection Agency. Any opinions, findings, conclusions, or recommendations are thos e of the authors and do not necessarily
reflect the views of EPA. This document is being circulated to obtain review and
comment from the Clean Air Scientific Advisory Committee (CASAC) and the general public. Comments on this draft document should be addressed to Dr. Anne Rea, U.S. Environmental Protection Agency, Office of Air Quality Planning and Standards, C539-02, Research Triangle Park, North Carolina 27711 (email: rea.anne@epa.gov).Chapter 7 – Synthesis and Integr ation of Case Study Results

DRAFT 7-1 August 2008 

## 7. SYNTHESIS AND INTEGRATION OF CASE STUDY RESULTS 1

The final version of this ch apter will synthesize the result s of the current conditions 2 assessments for each of the targeted ecosystem ef fect areas (i.e., aquatic acidification, terrestrial 3 acidification, aquatic nutrient enrichment, and terrestrial nutrient enrichment). Further, we will 4 use these summaries of ecological effects and their associated ecological indicators to 5 characterize the relationship between levels of an ecological indi cator and the associated degree 6 of ecologically adverse effects. This characterization of adversity can then be used to inform 7 policy judgments on public welfare. Once analyses are completed, th e next draft of this chapter 8 will describe the range of effects associated wi th the current deposition of nitrogen and sulfur 9
and demonstrate the links between adverse ecol ogical effects and the level of associated 10 ecological indicators. These relationships will th en be used to inform alternative standards by 11 translating adverse effects levels into levels of ecological indi cators that can be linked through 12 atmospheric deposition to ambient air concentrations of NO x and SO x . 13 Conducting a joint risk assessment for two criter ia pollutants that impact multiple (and in 14 some cases overlapping) ecological processes me ans that disparities exist in the level of 15 information available for the targeted ecosystem e ffect areas highlighted in this review. This is 16 reflected in the nature of the science, data , models, and time available for the case study 17 analyses. However, a concise synthesis of the information from the case studies, including an 18
assessment of the similarities and differences in the level of scientific data available to support 19 the risk assessments for each effect area, will be useful to inform the standard setting process. 20 The case study synthesis is intende d to provide an integrated, wei ght-of-evidence approach that 21 will consider the combined information from both evidence-based (i.e., relying primarily on 22
reference to the ISA review of published literature on the effects of the deposition of NO x and 23
SO x ) and empirical approaches (based on results of applications of data, models, and information 24 from the published literature), a nd their associated uncertainties, to explain how the risk and 25 exposure assessment can inform decision makers about what levels of protection might be 26 appropriate to protect public we lfare from adverse impacts on ecosystems. This synthesis will 27 look at the ecological indicators se lected for each of the targeted effect area and examine the 28 commonalities and differences among them. This information can then be used to inform a 29 policy judgment on a known or antici pated adverse public welfare effect. 30Chapter 7 – Synthesis and Integr ation of Case Study Results

DRAFT 7-2 August 2008 This chapter summarizes the current progress of the case studies (Section 7.1), begins to 1
discuss how the results of the case studies can be used to inform potential levels of protection 2 against ecosystem effects (Section 7.2), and prelim inarily points to how this might be used in 3 designing a meaningful secondary standard or standards for ambient concentrations of NO x and 4
SO x (Section 7.3). In this draft, we focus on the ecological effects of a quatic acidification, using 5 surface water ANC as an ecological indicator. As our analyses progress, we will include 6 additional ecological indicators and explain how they may be used to characterize adverse 7 ecological effects. 8



### 7.1 SUMMARY OF CASE STUDY ANALYSES 9

Table 7.1-1 summarizes the ecosystem characteristic s, ecological indicators, ecological 10 effects, and ecosystem services evaluated in th e case study analyses presented in Chapters 4 and 11
5. Note: currently these results reflect the presen t status of the first draft risk and exposure 12 assessment and will be updated for the sec ond draft as the analyses progress. 13

7.1.1 Aquatic Acidification 14 Acidic deposition has altered major biogeoche mical processes in the United States by 15 increasing the sulfur and n itrogen content of soils. This increase accelerates SO 4 2- and NO 3 ! 16 leaching from soil to drainage water, depletes base cations (namely Ca 2+ and Mg 2+ ) from soils, 17 and increases the mobility of toxic aluminum (A l) (see ISA Section 3.2) (U.S. EPA, 2008). Some 18 ecosystems and some areas of the United States ar e more sensitive than others to the effects of 19 acidic nitrogen and sulfur depos ition, particularly streams a nd lakes with naturally low ANC. 20 The impact of acidic deposition on aquatic sy stems is controlled by several environmental 21 factors, such as geology, surface water flow, so il depth, and weathering rates, all of which 22 contribute to the ability of an ecosystem to ne utralize the additional acidi c deposition and prevent 23 the lowering of ANC. The Adirondack Mountains of New York and the Blue Ridge Mountains 24 in the Shenandoah National Park of Virginia were selected for the eval uation of impacts from 25 acidic deposition of nitrogen and sulfur. These re gions receive relatively high levels of acidic 26 deposition and are known to be predisposed to ac idification, and extensive research, data, and 27
models exist for these areas. In ad dition, the results will likely be relevant in characterizing risks 28 to larger areas of the Appalach ian Mountains and the Northeast. 29Chapter 7 – Synthesis and Integr ation of Case Study Results

DRAFT 7-3 August 2008 ANC is a key indicator of acidification w ith relevance to terrestrial and aquatic 1 ecosystems. ANC is useful because it integrates the overall acid-base status of a lake or stream 2 and reflects how aquatic ecosystems respond to acidic deposition over time. There is also a 3 relationship between ANC and the surface water constituents that directly contribute to or 4 ameliorate acidity-related stress, in particul ar, concentrations of hydrogen ion (as pH), Ca 2+ , and 5 aluminum. Moreover, low pH surface waters leach aluminum from soils, which is quite lethal to 6 fish and other aquatic organisms. In aquatic syst ems, there is a direct relationship between ANC 7 and fish and phyto-zooplankton di versity and abundance (see Sec tion 7.3). These relationships 8 also impact several ecosystem services, such as fishing, recreation, f ood, natural habitat, and 9
tourism. 10

7.1.2 Terrestrial Acidification 11 Terrestrial acidification occurs where inorga nic and mineral acids ar e added to the soil. 12 This occurs as a result of both natural bioge ochemical processes and acidic deposition of 13 nitrogen and sulfur. Biological effects of acidification on terrest rial ecosystems are generally 14
attributable to aluminum toxic ity and decreased ability of plan t roots to take up base cations. 15 Acidic deposition increases concentrations of sulfur and nitrogen in soil, which accelerates 16 leaching of SO 4 2- and NO 3 - from soil to drainage water. In the absence of anthropogenic sulfur 17 and nitrogen deposition, the limited mobility of anions in the soil controls the rate of base cation 18 leaching. Atmospheric deposition of sulfur and nitrogen provides anions that are more mobile in the 19 soil environment than naturally occurring anions in the soil; these mineral acid anions can accelerate 20 natural rates of base-cation leaching, particularly Ca 2+ and Mg 2+ , leading to mobilization of inorganic 21 aluminum, which is toxic to tree roots. Acidic deposition can also a ffect terrestrial ecosystems by 22 causing direct impacts on plant foliage. 23 The regions considered to be most sensitive to terrestrial acidification effects due to 24 acidic deposition are forested ecosystems of the Adirondack Mountains of New York, Green 25 Mountains of Vermont, White Mountains of New Hampshire, Allegheny Plateau of 26 Pennsylvania, and high-elevation forests in the s outhern Appalachians. The Kane Forest in the 27 Allegheny Plateau of Pennsylvania and the Hubba rd Brook Experimental Forest in the White 28 Mountains of New Hampshire were selected to evaluate the e ffects of acidic deposition on the 29 sugar maple and red spruce due to known sensitiv ity of these species in these areas and the 30 availability of data. The ecosystem services aff ected include food, natural habitat, and tourism. 31Chapter 7 – Synthesis and Integr ation of Case Study Results

DRAFT 7-4 August 2008 For our case study areas, we anticipate that s ugar maple and red spruce abundance and growth 1 (i.e., crown vigor, biomass, and ge ographic extent) will be quantita tively linked to acidification 2 symptoms (specifically the calcium to aluminum ratio in the soil) through U.S. Forest Service 3 Forest Inventory and Analysis (F IA) database analyses and the an alysis of estimated sales of 4 maple sugar products for the second draft risk and exposure assessment. 1 5

7.1.3 Aquatic Nutrient Enrichment 6 Nitrogen is an essential nutrien t for estuarine and marine ecosystem fertility and is often 7 the limiting nutrient in aquatic ecosystems. Exce ssive nitrogen contributions can cause habitat 8 degradation, algal blooms, t oxicity, hypoxia (i.e., reduced dissolved oxygen), anoxia (i.e., 9 absence of dissolved oxygen), fish kills, a nd decreases in biodiversity. A recent national 10 assessment of estuary trophic conditions used five biological indicators to assess nutrient 11 enrichment: chlorophyll a , macroalgae, dissolved oxygen, nui sance/toxic algal blooms, and 12
SAV. This assessment is discusse d in detail in Attachment 5. 13 The Chesapeake Bay and Pamlico Sound were chosen as case study areas based on the 14 availability of atmospheric de position data, water quality mode ling data that accounted for 15 atmospheric deposition, a large main stem river, and scalability. Because the Chesapeake Bay 16 and Pamlico Sound are fed by multiple river system s, the case study areas were scaled to one 17 main stem river for each system: the Potomac River Basin for the Chesapeake Bay and the Neuse 18 River Basin for the Pamlico Sound. A eutrophicati on index, developed by National Oceanic and 19 Atmospheric Administration’s (NOAA’s) ) Na tional Estuarine Eutrophication Assessment 20 (NEEA) was used in this analysis, which estimates the likelihood that the estuary is experiencing 21 eutrophication or will experience eutrophication in the future. The endpoints of interest are fish 22 populations; water quality; and habitat quality an d their related ecosyst em services, including 23 fisheries (e.g., closings, decreased species richness) , recreation, and tourism. 24

7.1.4 Terrestrial Nutrient Enrichment 25 Terrestrial ecosystems typically respond to total reactive nitrogen loading and, with few 26 exceptions, have limited differential responses to the specific chemical species of deposited 27 nitrogen. Nitrogen compounds that are contained in atmospheric deposition, and that influence 28

1 The FIA can be found at http://fia.fs.fed.us/.Chapter 7 – Synthesis and Integr ation of Case Study Results

DRAFT 7-5 August 2008 nutrient dynamics, include multiple oxidized an d reduced forms of inorganic nitrogen. Excess 1 nitrogen in ecosystems causes inadvertent fertil ization of trees and gras slands, creating unnatural 2 growth rates in some species and changing comp etitive interactions am ong species, and nutrient 3 imbalances. These impacts ultimately decrease ec osystem health and biodiversity. Forest growth 4 enhancement can potentially exacerbate other nut rient deficiencies, such as Ca, Mg, or K, 5 thereby causing problems with forest health. En hanced growth generally occurs above ground 6 level. This can cause changes in the shoot-to-root ratio, causi ng decreased resistance to 7 environmental stressors, such as drought. In conifer forests, multiple long-term experiments (see 8 Attachment 6) have demonstrated transient growth increases (generally at deposition rates lower 9 than 10 kg N/ha/yr) followed by increased mortality, es pecially at higher rate s of fertilization. In 10 the western United States, atmos pheric nitrogen deposition has b een shown to cause increased 11 litter accumulation and carbon storage in abovegr ound woody biomass, which, in turn, may lead 12 to increased susceptibility to more severe fire s. Grassland communities that are adapted to low 13 nutrient supply can exhibit substantial sensitivity to nutrient enrichment effects of nitrogen 14 deposition. 15 Two of the primary indicators of nitrogen enrichment in fo rested watersheds are the 16 leaching of NO 3 - in soil drainage waters and the export of NO 3 -

  in stream water, especially during 17 the growing season. Low C:N ratios in soils are al so commonly related to increased nitrification, 18 potential increases in soil acidity, and releases of nitrate to receivi ng waters; however, these 19 measurements are not always widely available. 20
California’s CSS and mixed coni fer forest communities were selected for an initial case 21 study analysis. These areas were selected based on the availability of atmospheric ambient and 22 deposition data; the availability of digitized da tasets of biotic communities; fire-prone areas; 23 sensitive, rare species; their representativeness of western United States ecosystems potentially 24 impacted by nitrogen deposition; and their scalability and generalization potential. 25
California’s CSS and mixed conifer forest s on the slopes of the San Bernardino and 26
Sierra Nevada Mountains have im portant recreational value, prot ect water resources, and provide 27 habitats for many other species. In the mixed conifer forest, lichen communities and nitrogen 28 saturation can provide a means to identify the effects of nitr ogen loadings. Both habitats provide 29 a number of ecological services, including biodivers ity, water quality, recreation, timber, and fire 30 hazard mitigation. Linkage methods from endpoint to services may include measurement of 31Chapter 7 – Synthesis and Integr ation of Case Study Results

DRAFT 7-6 August 2008 changes in biodiversity and abundance and distri bution of threatened and endangered species, 1 comparison of past and present photography, an d measurement of the distribution of soil 2 moisture with depth and possible nitrate leaching. 3

7.1.5 Sulfur and Mercury Methylation 4 The biogeochemical cycle of mercury, a well- known neurotoxin, is closely tied to the 5 sulfur cycle. Mercury is take n up by living organisms in the methylated form, which is easily 6 bioaccumulated in the food web. Sulfate-reducing b acteria in wetland and lake sediments play a 7 key role in mercury methyl ation. Changes in sulfate 2 deposition have result ed in changes in both 8 the rate of mercury methylation and the corr esponding mercury concentrations in fish. 9 Although sulfur deposition is important to me rcury methylation, seve ral other interrelated 10
factors seem to also be rela ted to mercury uptake, including low lake water pH, dissolved 11 organic carbon, suspended particul ate matter concentrations in th e water column, temperature, 12 and dissolved oxygen. In addition, the proportion of upland to wetland land area within a 13 watershed, as well as wetland type and annua l water yield, appear to be important. 14 The major ecosystem services potentially impacted by sulfur-mediated mercury 15 methylation are provisioning and cultural services. Fishing and shellfishing can involve both 16 commercial operations and sports fishing, whic h provide food for human populations. For some 17 socioeconomic groups (especially groups with low in comes), fishing is a subsistence activity that 18 makes a very significant contribution to househol d food intake. Sport fishing often involves 19 important recreational services , and for many groups (e.g., Nativ e Americans, Alaska native 20 villagers), fishing and consuming local fish or sh ellfish is of cultural and spiritual significance. 21

2 EPA notes that in its review of the primary NAAQS fo r sulfur oxides, it has made the decision to examine all sulfate health effects as part of the review of the PM NAAQS (sulfates being one of the chief types of secondary PM). EPA believes it appropriate; however, to consider secondary sulfate effects as part of this review, rather than in the PM NAAQS review, because the e ffect at issue is related specifically to chemical composition. The fact that sulfate is deposited as a particle is not re levant to its contribution to mercury methylation.Chapter 7 – Synthesis and Integration of Case Study ResultsDRAFT 7-7 August 2008Table 7.1-1. Summary of Characteristics, Ecological Indicators, Effects, and Ecosystem Services Impacted for Each Case StudyEvaluated in This ReviewTargetedEffect AreaCharacteristics ofSensitivity (VariableEcological Factors)Biological/ChemicalCharacteristicEcologicalIndicatorEcological Effects EcosystemServicesImpactedCase StudyLocationsDeposition Levels(NOxand/or SOx)Associated withObserved Effects(from ISA)AquaticAcidificationGeology, surfacewater flow, soildepth, weatheringrates[Al]pHANCSpeciesrichness,abundance,compositionANCSpecies losses of fish,phytoplankton,zooplankton; changedcommunitycomposition,ecosystem structure,and functionFisheries,recreation,tourismAdirondackMountains,NYBlue RidgeMountains,ShenandoahNational Park,VA15 kg SO4-2/ha yr10 kg NO3-/ha yrTerrestrialAcidificationGeology, surfacewater flow, soildepth, weatheringratesSoil base saturation[Al][Ca]C:N ratioCa:Al ratioTree healthRed spruce,sugar mapleANCCa:Al ratioDecreased treegrowthIncreasedsusceptibility tostress, episodicdieback; changedcommunitycomposition,ecosystem structure,and functionFood,naturalhabitat,tourismKane Forest(AlleghenyPlateau, PA)HubbardBrookExperimentalForest (WhiteMountains,NH)AquaticNutrientEnrichmentN-limited systems,presence of nitrogenin surface water,eutrophication status,nutrient criteriaChlorophylla,macroalgae,dissolved oxygen,nuisance/toxic algalblooms, submergedaquatic vegetation(SAV)Eutrophication Index(EI)Habitat degradation,algal blooms,toxicity, hypoxia,anoxia, fish kills,decreases inbiodiversityFishpopulations,waterquality,habitatqualityPotomac RiverBasin,ChesapeakeBayNeuse RiverBasin, PamlicoSoundChapter 7 – Synthesis and Integration of Case Study ResultsDRAFT 7-8 August 2008TargetedEffect AreaCharacteristics ofSensitivity (VariableEcological Factors)Biological/ChemicalCharacteristicEcologicalIndicatorEcological Effects EcosystemServicesImpactedCase StudyLocationsDeposition Levels(NOxand/or SOx)Associated withObserved Effects(from ISA)TerrestrialNutrientEnrichmentPresence ofacidophytic lichens,anthropogenic landcoverCation ExchangeCapacity, C:N ratios,Ca:Al ratios, NO3-leaching and exportSpeciescompositionSpecies changes,nitrogen enrichmentof soil, changes infire regime, changesin nutrient cyclingLoss ofhabitat, lossofbiodiversity,recreation,waterqualityCoastal sagescrub andmixed coniferforest (SanBernardinoand SierraNevadaMountainRanges, CA)10–15 kg N/ha yr3–8 kg/ha yr forlichen effectsSulfur andMercuryMethylationPotentialWetland type,presence of sulfatereducing bacteria,water pH, dissolvedorganic carbon,suspended particulatematterInteraction amongdissolved organiccarbon, temperature,anoxia, and sulfideLand cover,precipitationresponse, andlimnographyMeHgconcentrations in fishand shellfishNeurotoxic effects infish and throughoutfood webFishing,shellfishing,sportsfishing,food,recreation,biodiversityLittle RockLake, WI (ISAcase study)Chapter 7 – Synthesis and Integr ation of Case Study Results

DRAFT 7-9 August 2008 

### 7.2 INTEGRATING CASE STUDY RESULTS 1 The ISA and case study analyses have shown th at, in certain ecosys tems, welfare effects 2 are occurring under current am bient concentrations of NO x and SO x , even though nationwide 3 (with only a few exceptions) we are in attainment of the current primary and secondary NO 2 and 4
SO 2 NAAQS. The next step in the risk and exposure assessment process is to determine what 5
alternative levels of ambient NO x and SO x might allow (1) various degrees of recovery of 6 impacted systems, (2) acceleration of recovery of impacted systems, and (3) protection against 7 further damage, recognizing the variability in ecosystem sensitivities to nitrogen and sulfur 8 deposition. At this point in the risk and exposure assessment, we have begun to address many of 9 the policy-relevant questions gui ding this review. Our initial ca se study analyses have begun to 10 show the variability associated with these responses and the degree to which ecological 11
indicators are suited to address a particular type of eff ect (i.e., acidifica tion or nutrient 12 enrichment in terrestrial or aquatic ecosystems). 13 There are many factors that determine whethe r or not an ecological effect occurs in 14 response to ambient concentrations of NO x and SO x . These “variable factors” may refer to either 15 ecological factors or atmospheric factors, both of which influen ce deposition or exposure and the 16 associated ecological effects (i.e ., acidification or nutrient enrich ment). The degree to which any 17 given level of atmospheric deposit ion of nitrogen and sulfur imp acts ecosystems depends in large 18 part on a suite of variable ecological factors at the point of deposition, including geology, surface 19 water flow, soil depth, and weathering rates. The levels of these ecological factors in a given 20 location determine the ecological conditions govern ing sensitivity to atmospheric deposition of 21 nitrogen and sulfur. The linkages between a gi ven ecological indicato r and deposition, as 22 influenced by the variable ecological factors, are discussed in Chapter 8. 23 In the next draft of the risk and exposure assessment, we will look across the results of 24 the case studies to identify ecological indicators that provide the most comprehensive coverage 25 across sensitive ecosystems. These indicators wi ll form the basis for assessing links between 26 adverse effects and potentia l levels of an ecological indicator to protect against those effects. At 27 this point, the Aquatic Acidificati on Case Study analysis is the furt hest along and will be used to 28 illustrate how an ecological indicator of acidification (i.e., ANC) and the levels of ecological 29Chapter 7 – Synthesis and Integr ation of Case Study Results

DRAFT 7-10 August 2008 effects associated with changes in that indicator can be used to characterize adversity to public 1
welfare. 2



### 7.3 LINKAGES BETWEEN ECOLOGIC AL INDICATORS AND ADVERSE 3


### EFFECTS 4 Within the limitations of currently availabl e data and models, our case study analyses 5 indicate the degree, if any, to which adverse ecological effects ar e already occurring in specific 6 sensitive ecosystems at current levels of ambi ent nitrogen and sulfur deposition. Note that 7 ecological recovery may not immediately result from decreases in atmospheric deposition, but 8 may lag behind changes in atmospheric depositio n. Since current ecological conditions reflect 9 past deposition and responses to it, it is possible that an ecolo gical indicator such as ANC may 10 improve over time (years) under current deposition le vels. However, the recovery rate may be 11 affected by current levels of deposition, with lowe r levels of deposition l eading to a more rapid 12 recovery compared to current levels. 13

7.3.1 ANC: The Critical Chemical Parameter 14
Ambient air concentrations are linked to a dverse ecological effects via the process of 15 atmospheric deposition. For our purposes, it is necessary to identify a suitably generalized 16 ecological indicator for these e ffects because linking individual ecological endpoints directly to 17 atmospheric deposition is difficult. The relations hip between ecological e ffects associated with 18 nitrogen and sulfur deposition and a comm only measured and well understood ecological 19 indicator, ANC, is described below. In addi tion, we conceptually de scribe linking ecological 20 effects with potential measures of adversity to public welfare by ch aracterizing impacts on 21 ecosystem services. This information can then he lp to inform judgments about the extent to 22 which different levels of a particular ecol ogical indicator might affect public welfare. 23 The susceptibility of a catchment to terrestri al or aquatic acidifi cation by nitrogen and 24 sulfur deposition is controlled by the ability of the watershed to neutralize the addition of 25
nonmarine acid anion concentrations of SO 4 2- or NO 3 - to the catchment. The neutralizing ability 26
of a catchment is a function of the acid-base balance; the balance between the sum of base 27 cations (i.e., Ca* + Mg* + K* + Na*) i nputs minus the sum of acid anions (i.e., SO 4 2- or NO 3 - ) 28 inputs. In general, catchments with positive acid -base balance have base cations inputs that 29Chapter 7 – Synthesis and Integr ation of Case Study Results

DRAFT 7-11 August 2008 exceed acid anions inputs which “buffer” the watershed from becoming acidic. The ability of a 1 catchment to buffer acid anion deposition is kno wn as the “buffering capacity,” which is 2 controlled by a host of environmental factors w ithin the watershed. Such factors include base 3 cation leaching rate, base cation weathering rates, base saturation, uptake by vegetation, rate of 4 surface water flow, soil depth, bedrock, etc. The bu ffering capacity of a catchment is controlled 5 by the flux of base cations, which act to balan ce acid anion inputs. The effect of acid anion 6 deposition is to lower this acid-base balance, causing a decline in the ecological system’s 7 composition, structure, and the biological fitness of its organisms over time. 8 The ANC of surface waters is widely used as a chemical indicator of acidic conditions 9 because it has been found in many studies to be the best single indicator of the biological 10 response and health of aquatic co mmunities in acid-sensitive system s (Lien et al., 1992; Sullivan 11 et al., 2006). Logistic regression of species presence/absence data against ANC provides a 12 quantitative dose-response func tion, which indicates the probab ility of occurrence of an 13 organism for a given value of ANC. For exampl e, the number of fish species present in a 14 waterbody has been shown to be positively correlated with the ANC level in the water, with 15 higher values supporting a greater ric hness and diversity of fish species ( Figure 7.3-1 ). The 16 diversity and distribution of phyto-zooplankton co mmunities are also positively correlated with 17
ANC. 18Chapter 7 – Synthesis and Integr ation of Case Study Results

DRAFT 7-12 August 2008 1 Figure 7.3-1. Number of fish species per lake or st ream versus acidity status, expressed 2 as ANC in lakes in the Adirondacks of New York (Sullivan et al., 2006). Five classes: <0, 3 0–20, 20–50, 50–100 μeq/L, and >100 μeq/L describe the biological effects at a range of 4 ANC concentrations. See Table 7.3-1 for more details. 5
For freshwater systems, ANC levels can be grouped into five major classes: <0, 0–20, 6 20–50, 50–100 μeq/L, and >100 μeq/L with each range representing a probability of ecological 7 damage to the community (see Attachment 3). ANC values above 100 μeq/L are generally not 8 harmful (Figure 7.3-1) to biota. Below 100 μeq/L, it has been shown that fish fitness and 9 community diversity begin to dec line, but the overall health of the community remains high as 10 long as ANC concentrations do not go below 50 μeq/L. ANC concentrations below 50 μeq/L 11 result in negative effects on sensitive biota. Fr om 50 to 20 μeq/L, it has been shown that fish 12 diversity and the overall fitness (i.e., health and reproduction) of most aquatic organisms in the 13 waterbody are reduced. Below 20 μeq/L, all biot a exhibit some level of negative effects, 14 particularly because surface waters at this level are susceptible to episodic acidification and their 15 associated harmful effects. Fish and plankton diversity and the structure of the communities 16 continue to decline sharply to levels where acidophilic specie s begin to outnumber all other 17Chapter 7 – Synthesis and Integr ation of Case Study Results

DRAFT 7-13 August 2008 species. Below an ANC of 0 μeq/L, complete loss of fish populations and extremely low 1 diversity of planktonic communities occur. At these low levels, only acidophilic species are 2 present, but even their popul ation and community structure are sharply reduced. The five 3 categories of ANC and expected ec ological effects are described in Table 7.3-1 and are 4 supported by a large body of research complete d throughout the eastern Un ited States (Sullivan 5 et al., 2006). The connection be tween the ecological effects described in Table 7.3-1 and 6 ecosystem services that can be indicative of advers e impacts to public welfare will be explored in 7 the second draft risk and exposure assessment. Ho wever, examples of these ecosystem services 8 include recreational and subsistence fishing, recrea tional lake use, natural habitat provision, and 9 tourism. Potential values of these ecosystem services related to alternative levels of ANC will be 10 assessed in the second draft risk and exposure assessment. 11
Table 7.3-1. Aquatic Status Categories Category Label ANC Levels Ex pected Ecological Effects Acute Concern <0 micro equivalent per Liter (μeq/L) Complete loss of fish populati ons is expected. Planktonic communities have extremely low diversity and are dominated by acidophilic forms. The numbers of individuals in plankton species that are present are greatly reduced.

Severe Concern 0–20 μeq/L Highly sensitive to episodic acidification. During episodes of high acidic deposition, brook trou t populations may experience lethal effects. Diversity a nd distribution of zooplankton communities decline sharply. Elevated Concern 20–50 μeq/L Fish species richness is great ly reduced (i.e., more than half of expected species are missing). On average, brook trout populations experience sub-lethal effects, including loss of health and reproduction (fitness). Diversity and distribution of zooplankton communities decline. Moderate Concern 50–100 μeq/L Fish species richness begins to decline (i.e., sensitive species are lost from lakes). Brook trout populations ar e sensitive and variable, with possible sub-le thal effects. Diversity and distribution of zooplankton communities begin to decline as species that are sensitive to acidic deposition are affected. Low Concern >100 μeq/L Fish species richness may be unaffected. Reproducing brook trout populations are expected where habitat is suitable. Zooplankton communities are unaffected and exhibit expected diversity and distribution. Whether the biological community of a waterbod y is at risk from current deposition loads 12 depends on the ability of the watershed to ne utralize the incoming aci d anion deposition and 13Chapter 7 – Synthesis and Integr ation of Case Study Results

DRAFT 7-14 August 2008 maintain an acid-base balance or ANC level that protects the biological community. This linkage 1 between deposition and biological pr ocesses is derived from a seri es of empirical relationships, 2
that relate the sustainable buffe ring capacity (preindustrial) of a catchment to the maximum flux 3 of acidic deposition which will maintain a particul ar level of ANC concentration in the surface 4 water. ANC is operationally defined as the sum of base cations minus the sum of acid anions, 5 with the proportion of ions derivi ng from neutral sea-spray removed: 6
ANC = [BC] t * - [AA] t * (1) 7
where [BC] t * is the current, measured sum of nonmari ne base cation equiva lent concentrations 8 (i.e., Ca* + Mg* + K* + Na*) and [AA] t * is the sum of nonmarine acid anion equivalent 9 concentrations (i.e., SO 4 * + NO 3 *). It is assumed that all chlo ride is derived from marine 10 sources. The * denotes the nonmarine component, wh ereby the marine contribution of each ion is 11 subtracted as a proportion of measured chloride concentration from the known ratio of these ions 12 in seawater. The equation defining ANC forms th e basis of the linkage between deposition and 13 surface water ANC concentrations. This linkage is explored in greater detail in Chapter 8. 14Chapter 7 – Synthesis and Integr ation of Case Study Results

DRAFT 7-15 August 2008 

### 7.4 REFERENCES 1
Lien, L., G.G. Raddum, and A. Fjellheim. 1992. Critical Loads of Acidity to Freshwater: Fish 2
and Invertebrates. The Environmental Tolerance Levels Programme. Rep. No. 23/1992. 3 Norwegian Ministry of Environment, Oslo, Norway. 4

Sullivan, T.J., C.T. Driscoll, B.J. Cosby, I.J. Fernandez, A.T. Herlihy, J. Zhai, R. Stemberger, 5 K.U. Snyder, J.W. Sutherland, S.A. Nierzw icki-Bauer, C.W. Boylen, T.C. McDonnell, 6
and N.A. Nowicki. 2006. Assessment of the Extent to Wh ich Intensively-Studied Lakes 7 are Representative of th e Adirondack Mountain Region . Final report. Albany, NY: New 8 York State Energy Research and Developm ent Authority (NYSERDA). Available at 9 http://nysl.nysed.gov/uhtbin/cgisi rsi/Qcwd6NzFby/NYSL/138650099/8/4298474 10 (accessed November 1, 2007). 11

U.S. EPA (Environmental Protection Agency). 2008. Integrated Science Assessment for Oxides 12 of Nitrogen and Sulfur – Environmental Cr iteria (Second External Review Draft)

. 13 Prepared by U.S. Environmental Protection Agency, National Center for Environmental 14 Assessment-RTP Division, Office of Research and Development, Research Triangle 15 Park, NC. EPA/600/R-08-082. August. 16Chapter 7 – Synthesis and Integr ation of Case Study Results

DRAFT 7-16 August 2008 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 [This page intentionally left blank.] 24 25Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-1 August 2008 

## 8. CONSIDERATIONS IN THE STRUCTURE OF NO X 

## /SO X 1 

## SECONDARY STANDARD 2
The previous chapters have provided an understanding of the risks associated with 3 current deposition of nitrogen and sulfur associat ed with ambient atmospheric concentrations of 4
NO x and SO x . Given that information, the next step is to provide a framework for developing 5
standards for NO x and SO x that provide the appropriate level of protection agains t those risks. 6 The purposes of this chapter are 1) to propose one possible framework for a NO x and SO x 7 secondary standard(s), based on the linkage s that exist between ambient atmospheric 8
concentrations of NO x and SO x , atmospheric deposition of N a nd S, ecological effects and the 9 related ecological indicators and 2) to illustrate how such a st andard might be expressed for 10 aquatic acidity using data from lakes in the Adirondacks case study area. The framework is 11 designed to both meet statutory and legal requirements, and to reflect the complex nature of 12 atmospheric and ecological systems. 13 In this chapter, we discuss constructing a st andard based on achieving a uniform level of 14 ecosystem protection. We explor e a framework by which ambient at mospheric concentrations of 15
NOx and SOx can be translated into a measure of ecosystem effects, using the transformation 16 functions described below. To facilitate a more concrete understanding of this framework, we 17 illustrate how the atmospheric concentration standards can be calculated using the Adirondacks 18 case study example. 19 The remainder of this chapter is organized as follows. In Section 8.1, we discuss a 20 possible structure for a secondary NAAQS for NO x and SO x , building on the conceptual model 21 introduced in Chapter 1. In section 8.2, we pr esent a discussion of the statutory and legal 22 considerations that affect the design of an ecolo gically relevant ambient air quality standard for 23
NO x and SO x . Section 8.3 discusses the framework for structuring a standard which links 24 atmospheric concentrations of NO x and SO x to a specific ecological in dicator of effects (i.e., 25 ANC) through the deposition process, taking atmosphe ric and ecological factors into account. It 26 also builds upon the current conditions risk a nd exposure assessment presented in Chapters 3 27 through 6 as well as the linkages presented in Chap ter 7 to discuss the nature of a meaningful 28 standard in light of the wide range of ecol ogical effects resulting from deposition of NO x and 29
SO x . In so doing, Section 8.3 expands on the sta ndards framework through which the linkages 30Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-2 August 2008 between ambient atmospheric concentrations of NOx and SOx, deposition of N and S, and 1 ecological indicators can be descri bed. Section 8.4, which is being developed for inclusion in the 2 next draft, will describe an illustrative example of how ambient atmospheric concentrations for 3 NOx and SOx can be compared to a desired level of ecological protection, as expressed using the 4 ecological indicator ANC. This example will al so show how observed (or modeled) ambient 5 atmospheric concentrations of NOx and SOx can be transformed into ANC values for 6 comparison against the level of a st andard defined in terms of a valu e of the ecological indicator. 7 This example will also provide step-by-step calculations to show how different combinations of 8 atmospheric concentrations of NO x and SO x can achieve the same level of ANC, based on the 9 deposition transformation functions and ecological e ffect functions that ar e described in Section 10 8.3. In each case, the example calculations will show clearly how the st andards framework can 11 incorporate 1) the influence of atmospheric fact ors that affect depositi on, 2) the influence of 12 ecological factors that affect ANC, and 3) the influence of reduced forms of nitrogen (ammonia 13 and ammonium). 14
15



### 8.1 POSSIBLE STRUCTURE OF A SECONDARY NAAQS 16 Throughout this assessment, we have recogni zed there are importa nt linkages between 17 ambient nitrogen and sulfur and their depositiona l effects on the environment. Our approach to 18 this joint, multi-pollutant review of two seconda ry standards has been to evaluate the science 19 driving welfare effects and consider what is needed to develop a secondary standard or suite of 20 standards that appropriately pr otects public welfare. The struct ure of the current secondary 21
standards for NOx and SOx is based on the ambien t concentration of a single indicator that is 22 specified as a nitrogen- or su lfur-containing compound (i.e., NO 2 and SO 2 , respectively) where 23 the level of the indicator does not necessarily correlate well with the ecological effects of 24 concern (USEPA, 2008). As discussed in Chapter 7, we recognize that the scientific basis that 25 forms our understanding of the re lationship between the criteria air pollutants, NOx and SOx and 26 their associated public welfare effects supports the use of a more complex structure for the 27 secondary standard(s). The inherently co mplex and variable linkages between ambient 28 concentrations of NOx and SOx, their deposited forms of nitrogen and sulfur, and the ecological 29Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-3 August 2008 responses that are associated w ith public welfare eff ects call for a more complex standard design 1 that reflects these linkages. 2
A critical difference between the current ri sk and exposure assessment and those from 3 previous reviews is that in the current assessment, exposures and ecological effects are related to 4 current levels of nitrogen and sulfur deposition in selected case study locat ions, rather than just 5
the ambient concentrations of NO x and SO x in those locations. Based on the recognized linkages 6
between ambient concentrations of NO x and SO x and deposition of nitrogen and sulfur (see 7 Chapter 3), as well as the linkage s between deposition and ecosystem effects, a specified level of 8 ecosystem effect will require limits on ambient concentrations of NO x and SO x , all other factors 9 being held constant. A standard that is expressed in terms of atmospheric concentrations of NO x 10
and SO x will require a process to translate ambient concentrati ons through the deposition process 11 into ecological effects, to determine if the specifie d level of ecosystem effect is achieved This 12 atmospheric concentration-based standard will exp licitly identify the role of modifying factors in 13 affecting the levels of NO x and SO x concentrations that would be necessary to achieve the 14
standard. 15 A possible structure for such standards is shown below in Figure 8.1-1, which builds 16 upon the conceptual model presented in Chapter 1. Figure 8.1-1 represents the pathway from 17 atmospheric concentrations of NOx and SOx to an ecological effect (aquatic acidification) as 18 represented by an ecological indicator (ANC), th rough the processes of deposition, with each of 19 the steps along the pathway expanded to reflect the example of aquatic acidification using the 20 ecological indicator ANC. Suites of variable factors reflecting both ecological and atmospheric 21 complexity modify the pathways. Critical to developing an ecologically appropriate secondary 22
standard for NO x and SO x is understanding that the different levels of the ecological indicator 23 can be associated with different levels of protec tion against adverse public welfare effects. The 24 ambient standard is then specified as the level(s) of ambient NO x and SO x such that the 25 calculated ecological indicator meets the leve l of protection of pub lic welfare set by the 26 Administrator. Thus, the standard is expressed in terms of th e relationship between ambient NO x 27
and SO x and the ecological indicator, linked th rough the deposition of N and S. 28Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-4 August 2008 1 Figure 8.1-1. Possible structure of a s econdary NAAQS for NOx and SOx based 2 on the ecological indicator ANC for the effect of acidification. 3
4 The variable atmospheric factors which infl uence atmospheric conditions affecting the 5 transformation, transport and depositi on of ambient concentrations of NO x and SO x include 6 relative humidity, precipitation, wind conditions, and chemical reactivity. Deposition is also 7 affected by seasonally dependent landscape factors such as leaf surface ar ea. Since deposition is 8 the driver for many of the associ ated ecological effects due to NO x and SO x , such understanding 9 of the factors that link atmosphe ric concentrations with deposition is critical. In addition, the 10
relative contribution of NO x to total deposition of nitrogen is affected by the ambient 11
concentrations of ammonia and ammonium. Th e linkages between atmospheric concentrations 12 and deposition as influenced by the variable at mospheric and landscape factors are discussed in 13
Section 8.3.1. 14Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-5 August 2008 The degree to which any given level of depositio n of nitrogen and sulfur containing species 1 impacts ecosystems depends in large part on a suite of variable ecological f actors at the point of 2 deposition, including soil base saturation, soil th ickness, runoff rate, land use conditions, nutrient 3 cycling, and non-atmospheric loadings of sulfur and nitrogen. The particular levels of these 4 factors that occur at any location form the ecolo gical baseline against which ecological effects 5 from deposition are evaluated. The linkages be tween deposition and ecological effect as 6 influenced by the variable ecological factors are discussed in Section 8.3.2. 7
The relationships illustrated in Figure 8.1-1 in concept can be operated in either direction, 8 from known or assumed ambient atmospheric concen trations to the resul ting ecological effects, 9 or from desired ecological conditions in an area of interest to ambient concentrations that would 10 achieve them. In either approach, there may not be a unique combination of ambient NOx and 11 SOx that achieves a specified level of eco logical protection, becau se of the multiple 12
combinations of NO x and SO x concentrations that can yiel d the same level of ecological 13 indicator. This is particularly true for the case of aquatic acidity, but may not hold true for other 14 ecological endpoints. Whether precipitation occurs or not also greatly af fects the relationship 15 between ambient atmospheri c concentrations of NO x and SO x and deposition of N and S, so 16 knowledge or an assumption of precipitation is n eeded to complete the quantitative trail from 17 ambient concentrations to ecological effects, or the reverse. Finally, concentrations and 18 deposition of reduced nitrogen species influence ecological conditions, but these species are not 19 currently encompassed by the listed criteria pollutant NO x . While reduced species of nitrogen do 20 contribute to the overall loadings of N, as will be demonstrated in a future version of Section 8.4, 21 it may be possible to focus solely on NO x contributions to loadings while recognizing that there 22 are impacts from reduced forms of nitrog en that must be taken into account. 23



### 8.2 STATUTORY AND LEGAL CONSIDERATIONS IN THE 24 

### STRUCTURE OF THE STANDARD 25 Inherent in the approach de scribed above are various lega l considerations as to the 26 requirements of the CAA and EPA’s ability to implement various types of standards under the 27 Act. This section will begin to discuss the le gal framework for considering appropriate air 28 quality indicator(s), averaging time(s), form(s), and level(s) of standard(s) to protect against 29 known or anticipated adverse imp acts on public welfare. As noted above, the structure of such a 30Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-6 August 2008 standard would likely be comple x, reflecting the variable factor s defining the atmospheric and 1
environmental conditions that affect deposition of NO x and SO x , and those variable ecological 2
factors that affect ecosystem responses. 3 In setting NAAQS to protect public health a nd welfare, EPA has historically established 4 standards which require the comparison of locally monitored ambient con centrations of an air 5 pollutant indicator against a specified numeri cal level of atmospheri c concentration of a 6 pollutant, using a specified aver aging time and statistical form. For example, the current 7 secondary standard for NO x uses ambient concentrations of NO2 as the indicator, and attainment 8 is determined by comparing the annual arithmetic mean of measured ambient concentrations, for 9 a calendar year, against the level of 0.053 ppm. As more is learned about the effects of pollutants 10 such as NOx and SOx on the environment, such a relatively simple framework may not be the 11 most appropriate way to protec t sensitive ecosystems from effects associated with ambient 12 concentrations of the air pollutants. EPA is consid ering in this review of the secondary standard 13
for NO x and SO x a conceptual framework for evaluating the environmental effect of ambient 14
concentrations of NO x and SO x that takes into account variable factors, such as atmospheric and 15 ecosystem conditions that modify the amounts a nd effects of deposited N and S on ecosystems. 16 EPA is also considering whether the secondary standard can and should be structured to better 17 take this framework into account, as a way to ap propriately protect the pu blic welfare from the 18 effects associated with the presence of these pollutants in the ambient air. 19 As discussed previously, this could involve a standard with an ecol ogical indicator which 20
can be linked to ambient concentrations of SO x and NO x . The standard could include a nationally 21 uniform level for a particular eco logical indicator, for example a specified level of ANC that 22 would be determined by the Administrator to repr esent the appropriate le vel of protection. This 23
framework would also take into account other variable s, both atmospheric and ecological that 24 may modify the ecological effects. Complia nce would be determined by measuring NO x and SO x 25 concentrations, using these measurements and th e various transformation functions to calculate 26 an ANC level, and comparing the calculated level of ANC to the level of ANC set by the 27 Administrator, including averaging time and form. Since the effect of a certain ambient 28
concentration of SO x and NO x on ANC could vary depending on the other variable factors, 29
whether a certain ambient concentration of SO x and NO x would attain the standard would depend 30 in part on these other variables, which c ould vary geographically and temporally. 31Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-7 August 2008 One important issue in evaluating such an a pproach is EPA’s authority to structure a 1 NAAQS in such a way. EPA’s review of a NAAQ S starts with its review of the air quality 2 criteria for the pollutant. Under Section 109(d )(1) of the CAA, EPA must undertake a thorough 3 review of the air quality criteria for the polluta nt at issue as part of reviewing a secondary 4 NAAQS to determine whether a current standard is requisite to protect the public welfare. Under 5 Section 108, air quality criteria ar e to “reflect the late st scientific knowledge useful in indicating 6
the kind and extent of all identifiable effects” a ssociated with the presence of the pollutant in the 7
ambient air. It is clear from the language of the CAA that where the state of the science provides 8 a basis for considering such effects as releva nt, the review of air quality criteria should 9 encompass a broad analysis of “any” anticipated a dverse effects, as well as the ways in which 10 variable conditions such as atmospheric condit ions or ecological conditions may influence the 11 effect of a pollutant and the ways in which othe r air pollutants may intera ct with the criteria 12 pollutant to produce adverse effects. Specifical ly, Section 108(a)(2) of the CAA provides that: 13 The criteria for an air pollutant to the exte nt practicable shall include information on the 14 following: 15
1. Those variable factors (including atmospheri c conditions) which of themselves or in 16 combination with other factors may alter th e effects on public health or welfare of 17 such air pollutants; 18 2. The types of air pollutants which, when pres ent in the atmosphere, may interact with 19 such pollutants to produce an adverse effect on public he alth or welfare; and 20 3. Any known or anticipated a dverse effects on welfare. 21 22 Taking into account this extensiv e review of the air quality criteria for an air pollutant, 23 the Administrator is required to review and to re vise, as appropriate, the secondary standard. The 24 secondary standard is to “specify a level of air quality which, in the judgment of the 25 Administrator, based on such criteria , is requisite to protect pub lic welfare from any known or 26 anticipated adverse effects associated with the pr esence of such air pollutant in the ambient air.” 27 CAA § 109(b) (emphasis added). “Welfare,” in tu rn, is defined to include a broad array of 28 effects, including effects on soil, water, crops , vegetation, and manmade materials, “whether 29 caused by transformation, conversion, or combina tion with other air pol lutants.” CAA § 302(h). 30 As with the description of th e air quality criteria in Section 108, the CAA uses expansive 31 language in describing the range of effects that EPA should take into account in setting a 32 standard that is requisite to protect public welfare. 33Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-8 August 2008 Thus, Section 108 and the definition of welf are require that EPA analyze and evaluate 1
scientific information on the interaction between ambient concentrations of a pollutant and other 2 variable factors, atmospheri c and ecological, that affect the air pollutants impact on the 3 environment. Section 109(b) makes it clear that EPA must base the standard on this scientific 4 information but does not specify how EPA is to inco rporate this information into the structure of 5 the standard beyond the requirement that any seconda ry NAAQS be a level of air quality that is 6 requisite to protect public welfar e. Accordingly, while Sections 108 and 109(b) make it clear that 7
EPA must take this kind of information in to account when setting the NAAQS, specific 8 directions are not provided on how to accomplish this. 9 It is also important to keep in mind the mean ing of the term “requisite,” as interpreted by 10 EPA and the Courts. In considering the meaning of the term “requisite” in the context of the 11 primary standards, the Supreme Court has agreed with EPA that such a standard is one that is 12 “sufficient, but not more than n ecessary” to protect public health. Whitman v. American 13 Trucking, 531 U.S. 457, 473 (2001). In addition, while Section 109(b) requires that EPA 14 establish “national” standards, it does not define this requireme nt in any greater detail than 15
described above. 16
While EPA has most often made decision based on the results of direct exposure to an air 17 pollutant in the ambient air in assessing eff ects on public health and welfare in establishing 18 NAAQS, such as the health effect s on humans when breathing in an air pollutant or the effects 19 on vegetation through the uptake of air pollutants through leaves, EPA has also considered, 20 where appropriate, the effects of exposure to air pollutants through more indirect mechanisms. 21 For example, in 1978, EPA established a NAAQS for lead that included non -respirable particles 22 based on the conclusion that these particles co ntribute to blood lead levels once they are 23 deposited on the ground and then become a source of human lead exposur e through the ingestion 24 of lead-contaminated food or, particularly in children, the placing of hands and other 25 contaminated objects in the mouth. Lead Industries v. EPA , 647 F.2d 1130 (DC Cir. 1980). In 26 addition, EPA has previously considered acidif ication and nutrient enrichment concerns 27
associated with ambient concentrations of NO 2 and other nitrogen compounds, but determined 28
that the scientific and technical evidence that was available at the time did not provide an 29 adequate basis to set a separate secondary st andard to address such effects. 61 FR 52852, 52855 30 (October 8, 1996). The deposition of ambient NO x and SO x to terrestrial and aquatic 31Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-9 August 2008 environments can affect ecosystems through direct and indirect mechanisms, as discussed in 1 previous chapters. Given Congress’ instruction to set a standard th at “is requisite to protect the 2 public welfare from “ any known or anticipated adverse effects associated with the presence of 3 such air pollutant in the ambi ent air,” 42 U.S.C. § 109 (b)(2) (e mphasis added), this review 4 appropriately attempts to take into consider ation widely acknowledged effects, such as 5 acidification and nutrient enrichment, which ar e associated with the presence of ambient SO x and 6
NO x , as described in previous chapters. 7 The relationships between the impacts of ambi ent air concentrations on ecological effects 8 such as acidification and nutrient enrichment, however, are extremely complex. To take these 9 effects into account, including th e variability in effects resu lting from the same ambient 10
concentrations of SO x and NO x , EPA is developing a biologically relevant framework in this risk 11
assessment that relates ambient concentrations of an air pollutant to an ecological indicator, 12 taking into account relevant atmos pheric and ecological variables. In that context it is useful to 13 explore whether a standard struct ured to take these elements into account is an appropriate way 14 to provide protection that is su fficient but not more than necessa ry, as compared to a standard 15
that treats ambient concentrations the same i rrespective of these othe r biologically relevant 16 variables. The CAA, which requires EPA to esta blish “national” standards, does not specifically 17
address whether EPA can establish standards which could allow for varying levels of SO x and 18
NO x in different areas. However, such an approach may be necessary if EPA is to establish a 19 secondary standard that provides sufficient but not more than necessary to protect the public 20 welfare. Given the lack of specificity in Section 109(b), EPA believes it is appropriate to explore 21 the approach described in this chapter, with an aim of setting a national standard that provides 22 sufficient but not more than necessary protec tion throughout the United States by taking into 23 account the recognized variation of biologically relevant factors. 24



### 8.3 LINKAGES FOR STRUCTURING ECOLOGICALLY RELEVANT 25 

### STANDARDS 26 Consistent with the framework depicted in Figure 8.1-1, the followi ng sections provide a 27 discussion of both the linkages between ambien t air concentrations and resulting deposition 28 metrics, and between the deposition metric a nd the ecological indicator of concern. These 29 linkages are discussed conceptu ally in Section 8.3.1. Followi ng that discussion, Section 8.3.2 30Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-10 August 2008 describes the deposition transfor mation function that relates atmospheric concentrations of NOx 1 and SOx to nitrogen and sulfur deposition metr ics, while Section 8.3.3 describes the ecological 2 effect function that transforms the deposition me tric into the ecologica l indicator. Modifying 3
factors that alter the relations hip between ambient atmospheric concentrations of NOx and SOx 4 and depositional loads of nitrogen and sulfur, an d those that modify th e relationship between 5 deposition loads and the ecological indi cator are also discussed more fully. 6

8.3.1 A Conceptual Model of the Pa thway from Ambient Atmospheric 7 Concentrations of NOx and SOx to Ecological Effects 8 For this first draft risk and exposure assessm ent, we have focused on developing an air 9 quality indicator for protection against the adve rse effects of aquatic acidification. Figure 7.3-1 10 demonstrates the various levels of protection of fish populations associated with varying ANC 11 levels. Differing levels of an ecological indicator such as ANC will be associated with differing 12 levels of protection of ecosystem functions (e .g. fish populations, biolog ical diversity, fish 13 population health). Different ecological indicato rs may be necessary for protection against the 14
adverse effects of terrestrial acidification and aquatic and terrestrial nutrient enrichment. 15
Indicators for other welfare effects categories of concern will be expl ored in future drafts of this 16 risk and exposure assessment. 17 We have created a system of linked functions that translate an atmospheric indicator 18 (concentrations of NO x and SO x ) into an ecological indicator that expresses the potential for 19 deposition of nitrogen and sulfur to acidify an ecosystem (Figure 8.1-1). The first part of this 20 system is the deposition tran sformation function, which transforms ambient atmospheric 21 concentrations (the atmospheric i ndicators) into a deposition metric. In this particular example, 22 the deposition metric referred to in Figure 8.1-1. is the sum of atmospherically deposited 23 oxidized nitrogen and sulfur, weighted by their mola r ratio. The second part of this system of 24 functions is an ecological effect function which converts the deposition me tric to an ecological 25 indicator (ANC) by adjusting for levels of othe r ecological factors that contribute to ANC. 26 When the amounts of NOx and SOx in the ambient atmosphere can be used to derive a 27 deposition metric (via the deposi tion transformation function) that can then be used to derive a 28 level of ANC (through the ecological effect functi on) which falls within the range defined as 29 acceptable by the standard, then those levels of NOx and SOx will be considered to meet that 30Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-11 August 2008 standard of protection. The atmosp heric levels of NOx and SOx th at satisfy a particular level of 1 ecosystem protection are those levels that result in an amount of deposition that is less than the 2 amount of deposition that a given ecosystem can accept without degradation of ANC. This 3
latter amount is referred to as the maximum de position load and is the amount that solves a 4 mass-balance equation for that particular level of ANC. Mass-balance rela tionships developed in 5 the scientific literature provide the basis for es tablished equations that link levels of ANC to 6 levels of nitrogen and sulfur de position (given levels of other, non-deposition related parameters 7 such as weathering rates, base cation uptake rates). Quantitative examples of these relationships 8 will be provided in Section 8.4 of this chapte r in the next draft of this document. 9 The value of the deposition metric can be calcu lated in terms of measured (or modeled) 10
ambient NO x and SO x concentrations by using establis hed relationships between ambient 11
concentrations of NO x and SO x and nitrogen and sulfur deposition. As will be described in 12 Section 8.4, these relationships are derived from the physical processes th at govern transport and 13 deposition; they are species de pendent; they vary by particle mass, meteorology, compensation 14 point, terrain and land use type. 15 It is important to note for this type of ai r quality indicator that the same level of 16 protection from acidification can be obtained with different combinations of ambient NO x and 17
SO x concentrations and their associat ed levels of nitrogen and sulf ur deposition. As a result, the 18 ambient air quality standard for NO x and SO x may be expressed jointly, and the air quality 19 standard may be met by various combinations of NO x and SO x . 20 Determining the level of the air quality indicato r that is associated with a particular level 21 of protection against adverse effects of acidificat ion may include consideration of factors that 22 could potentially alter th e depositional relationship of nitrogen or sulfur to ANC. In addition, the 23 specific concentrations of ambient NO x and SO x that will result in a particular deposition metric 24 can vary based on factors that affect the re lationship between ambient concentrations and 25 deposition. These two considerations im ply that the levels of ambient NO x and SO x that would 26 be consistent with a particular level of AN C may differ according to the specific levels of 27 modifying factors in different locations and whether the ecosyst em is nitrogen and/or sulfur 28
limited or saturated. While the standard is set to achieve the same le vel of protection (ANC) 29 nationwide, the atmospheric concentrations of NOx and SOx and the related amounts of 30 deposition that can occur and stil l meet the standard may differ. For example, because of the 31Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-12 August 2008 differences in maximum depositional loads be tween watersheds with high proportions of 1 grasslands and those with high proportions of ever green forest ecosystems, for the same level of 2 ANC, there may be differences in the ambient c oncentrations of NOx and SOx that would meet 3 that ANC level. The response variable associated with these examples will vary across effects 4 categories, and might be based on soil base saturation or aluminum concentrations in soil water 5 or biodiversity impacts, rather than ANC. Specific modifying factors for the ambient 6 concentration to deposition relationship are discu ssed in Section 8.3.2 of this chapter. Specific 7 modifying factors for the deposition to acidifica tion relationship will be discussed in Section 8
8.3.3. The overall transformation function of the sta ndard will be national in scope as will be the 9 ecological indicator level, alt hough specific modifying factors with in the transformation function 10 and maximum depositional load may vary by loca tion. Therefore, while the Administrator may 11 choose the level of protection that is desired by specifying a particular limiting value of ANC 12 (ANC limit ), there are a series of transformations th at link that level of protection to ambient 13
concentrations of NO x and SO x . As a conceptual example, to express the previous discussion in 14 mathematical terms, consider the following: 15

I= ANC limit = g(·)-DL(S)-DL(N) (1) 16
where I is an indicator, g(·) is a series of non-atmospheri c, non-depositional environmental 17 factors (e.g., runoff, weathering rates, geology, etc.), ANC limit is a critical limit for ANC 18 necessary to provide a particular level of prot ections for biota in an ecosystem, DL(S) is the 19 implied maximum depositional load of S associat ed with that ANC limit, and DL(N) is the 20 implied maximum depositional load of N associated with that ANC limit. Note that DL(S) and 21 DL(N) are jointly determined by the equation g(·)-ANC limit =DL(S)+DL(N), and that there is no 22 unique solution for either one unles s the other is fixed. Further, we can express the deposition of 23 NOx and SOx through the following relationships: 24
25 S D =f S (NO x , SO x |Z) and N D =f N (NO x , SO x |Z) (2) 26 27 where S D is sulfur deposition, N D is nitrogen deposition, Z is a series of parameters affecting 28 deposition of ambient SO x and NO x (including NH 3 ), f S is a function relating S D to ambient 29
concentrations of NO x and SO x , given Z, and f N is a function relating N D to ambient 30
concentrations of NO x and SO x , given Z. 31Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-13 August 2008 It is clear that the indicator of NO x and SO x effects (I) is a function of deposition. 1 However, by substituting in the ambient to depos ition conversion functions in (2), the indicator I 2
can be made into an ambient indicator: 3
I= g(·)-S D (f S (NO x , SO x |Z)-N D (f N (NO x , SO x |Z). (3) 4

Not all values of S D and N D cause adverse welfare effects. The point where the levels of 5 deposition associated with the ambient NOx and SOx lead to an ANC that does not achieve the 6 specified ANC limit are when S D and N D are equal to or greater then the DL(S) and DL(N) for a 7 particular location. 8
I= g(·)-DL(f S (NO x , SO x |Z)-DL(f N (NO x , SO x |Z). (4) 9
Clearly, the establishment of th e functional forms for g(·), f S , and f N , as well as for the 10 values of the parameter series Z, will be critical. Note that variable factors in g(·) and Z can be in 11 many forms, some continuous, e.g. precipitation, and some discrete, e.g. type of bedrock. 12 However, continuous variables can be converted to discrete categories, which may be easier to 13 implement in the form of a standard. For example, precipitation could be expressed as inches per 14 year, or precipitation could be categorized as high, medium , and low, possibly based on 15 percentiles of the observed distri butions of precip itation values. 16

8.3.2 Linkages between Ambient Concentrations and Deposition 17 Defining the linkages between ambient concentrations of NO x and SO x and deposition of 18 N and S is a crucial piece of the framework as described in Section 8.1. Atmospheric deposition 19 of nitrogen from ambient concentrations of NO x (N NOx ) and atmospheric deposition of sulfur 20
from ambient concentrations of SO x (S SOx ) to a surface is a complex function of numerous 21 meteorological parameters and su rface properties. A tr ansformation function that incorporates 22 the nationwide variability of the important meteor ological and surface parame ters is expected to 23 be a useful approach to describing the depositi on process. The derivation of this transformation 24 function, which converts ambi ent concentrations of NO x and SO x to their associated depositions, 25
is described below. The deriva tion takes algorithms from the U.S. EPA-developed Community 26 Multi-scale Air Quality (CMAQ) modeling syst em (U.S. EPA, 1999) into consideration, and 27 CMAQ output is used to calculate key parameters in the resulting func tion. The use of CMAQ 28Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-14 August 2008 data is intended to be illustrative of the pr ocess. Specific models, data sets and methods 1 appropriate to use in other locations or on di fferent spatial/temporal scales may differ. 2

8.3.2.1 Data and Tools 3
8.3.2.1.1 Physics and Chemistry of Atmospheric Deposition 4
Atmospheric pollutants deposit onto land and water surfaces through at least two major 5
mechanisms: direct contact w ith the surface (dry deposition) , and transfer into liquid 6 precipitation (wet deposition). The magnitude of each deposition process is related to the 7 ambient concentration through the time-, location-, pro cess- and species-specific deposition 8 velocity (Seinfeld and Pandis, 1998): 9

Amb i Dry i Dry i C v Dep

⋅ = (19) 10

Amb i Wet i Wet i C v Dep

⋅ = (20) 11

where v Dry i and v Wet i are the dry and wet deposition velocities, Dep Dry i and Dep Wet i are the dry and 12 wet deposition fluxes, C Amb i is the ambient concentration, and the i subscript indicates the 13 pollutant species under study. The total deposition of each pollutant is 14

Wet i Dry i Tot i Dep Dep Dep + = (21) 15
Substituting Equations 19 and 20 into 21 yields 16

Amb i Wet i Amb i Dry i Tot i C v C v Dep

⋅ + ⋅ = (22) 17
The total deposition of sulfur a nd nitrogen would therefore be: 18 

# ∑

⋅ ⋅ + = + i Amb i i Wet i Dry i Tot N S C m v v Dep ) ( (23) 19

where m i is the molar ratio of the atom (sulfur or nitrogen) of interest to the i th pollutant. 20 Ambient sulfur- and nitrogen-cont aining pollutants include gases such as sulfur dioxide (SO 2 ), 21 ammonia (NH 3 ), various nitrogen oxides (NO, NO 2 , N 2 O 5 ), nitric acid (HNO 3 ), and peroxyacetyl 22
nitrate (PAN); and particulate species such as sulfate (SO 2  4 ), nitrate (NO  3 ), and ammonium (NH + 4 23 ). Oxides of nitrogen and oxi des of sulfur are currently regu lated by the NAAQS as defined in 24Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-15 August 2008 the CAA. These compounds include the sulfur -containing species above as well as the above 1 oxidized forms of nitrogen (referred to here as NO x ); ammonia and ammonium, which are 2 reduced forms of nitrogen, are not curre ntly included as re gulated pollutants. 3 For this section, Equation 5 is termed the deposition transformation function , and it 4 provides the functional form of the relationshi p between instantaneous atmospheric pollutant 5 concentrations and deposition flux es of sulfur and nitrogen. E quation 5 can also be thought to 6 represent the relationship between cumulative or av erage values of the variables over time and/or 7 space, but doing so raises issues of exactly how integration/averaging over time and/or space is 8 to be done. To some extent, the degree of tempor al and spatial averagi ng needed will depend on 9 the ability to measure or model c oncentrations at particular spa tial (e.g. at specific water bodies) 10 and temporal (e.g. hourly or daily) scales. Ho wever, the need for temporal or spatially 11 aggregated values may also be driven by the fo rm of the ecological indicator, which may be 12 annual or seasonal, and may reflect local ecologi cal attributes. The remainder of Section 8.3.2 13 discusses this and other policy-re levant scientific and mathemati cal issues and proposes methods 14 for determining parameter values (i.e. v i dry and v i wet ) for input to Equation 23; results of example 15
calculations will be shown in Section 8.4. 16
8.3.2.1.2 Air Quality Simulation Models 17
Deposition velocities necessary for employing the deposition transf ormation function can 18 be obtained either through mode ls or direct measurements, a lthough the logistics of accurate 19 measurements are currently prohibitive to depl oying a measurement network. For purposes of 20 the current discussion and accompanying future example calculations (Section 8.4), deposition 21 velocities will be obtained from CMAQ output, and research will be conducted to assess if this is 22 the optimal source of these parameters for future regulatory analyses. 23 CMAQ determines the deposition fluxes of the suite of modeled air pollutants at each 24 time step and location in the us er-specified spatial grid. Depo sition velocities can be found in 25 CMAQ output in the following ways: 26
1. v dry values of gaseous pollutants are calc ulated in the CMAQ weather module called 27 the Meteorology-Chemistry Interface Pro cessor (MCIP) through a complex function 28 of meteorological parameters (e.g. temperat ure, relative humidity ) and properties of 29 the geographic surface (e.g. leaf area index, surface wetness) 30Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-16 August 2008 2. v dry values for particulate pollutants are cal culated in the aerosol module of CMAQ, 1 which, in addition to the parameters n eeded for the gaseous calculations, also 2 accounts for properties of the aerosol size distribution 3 3. Equation 20 is not directly expressed in CMAQ source code: the cloud processing 4 module of CMAQ performs simulations of mass transfer into cloud droplets and 5 aqueous chemistry to incorporate pollutants into rainwater, all of which is 6 conceptually contained in the v wet parameter. Explicit v wet values can and will be 7 calculated in a post-processing step by applying Equation 20 to the wet depositions 8 and ambient concentrations output by the model. 9
Due to lack of direct measurements, no performance evaluations of CMAQ’s dry 10 deposition calculations can be found; however, the current state of MC IP is the product of 11 development work that has been based on peer-r eviewed literature from the past two decades 12 (U.S. EPA, 1999) and is considered to be EPA’s best estimate of dry deposition velocities. Some 13 bias has been found between CMAQ’s wet depos ition predictions and measured values (Morris 14 et al., 2005); recent analyses suggest that poor simulation of precipitation could be responsible 15 for this (Davis and Swall, 2006), which should be considered if v i wet values are applied to 16 ambient measurements (this will be discussed in Section 8.4). Although the model is continually 17 undergoing improvement, CMAQ is EPA’s state-of-the-science computational framework for 18 calculating deposition velocities, and was therefore the first c hoice for exploring inputs to 19 Equation 23. Gaseous v dry values are directly output by MCIP, whereas aerosol v dry values and all 20 species’ v wet values currently need to be calculate d from standard CMAQ output. In all cases, 21 data aggregation issues need to be carefully considered: while MCIP and CMAQ give hourly 22 output values, the particular averaging time relevant for ecological processes for NO x and/or SO x 23 is likely to be the same as the time resolution of the ecological indicator of choice (which is 24 likely to be seasonal or a nnual). As such, values of v dry and v wet for each species should either be 25 representative of the desired av eraging period, or aggregation of resulting depositions after 26 applying hourly deposition velocitie s to concentrations will need to be performed. It is proposed 27 that the spatial scale of the deposition velocities used here reflect that of future analyses, e.g. be 28 relevant at scales that may ra nge from individual water bodies up to regional or national scales, 29 as discussed in Chapters 4 and 5. Specific me thods for performing calculations are discussed 30
below. 31Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-17 August 2008 8.3.2.2 Analytical Techniques 1
8.3.2.2.1 Hourly Deposition Velocities 2
As stated above, hourly v dry values of gaseous specie s are complex functions of 3 meteorology and land surface properties that are directly available in the outputs of MCIP. 4 Deposition velocities of aerosol species were ca lculated by applying Eq uation 19 to CMAQ grids 5

of aerosol C Amb i and Dep Dry i values to calculate hourly grids of v Dry i values for the example in a 6 future Section 8.4. Plans are underway to obtai n the size-specific aero sol deposition velocities 7 from future CMAQ model runs. This will provide v Dry i for fine as well as coarse mode aerosols. 8 Unlike dry deposition velocity, the wet depositio n velocity is independent of land surface 9 properties and can be effectivel y expressed in a simple form (Seinfeld and Pandis, 1998): 10

0 p w v r i Wet i ⋅ = (24) 11

Amb i Aq i r i C C w = (25) 12

where w r is the washout ratio, C Aq i is the concentration of species i in the rainwater, and p 0 is the 13 precipitation intensity. 14 If Equation 25 is substituted into 24, and E quation 24 is substituted into Equation 20, the 15 ambient concentration cancels out and wet depos ition can be expressed as a function of only 16 rainwater concentration and prec ipitation (both of which are routinely measured by the NADP 17 network). However, deposition can be expressed in terms of ambient concentrations since the 18 atmospheric indicators for the framework discus sed above are ambient air concentrations of NO x 19
and SO x . Calculating the transformation function will therefore use values of w r derived from 20 CMAQ output for input into Equation 24 (which is then input to Equation 5). Solving for v i wet in 21 Equations 2 and 6, equating them and rearranging these equations yields: 22

Amb i wet i r i C p

Dep w

⋅ = 0 (26) 23

which expresses the washout ratio in terms of CMAQ outputs. As mentioned above, simulated 24 precipitation is a source of e rror in CMAQ’s wet deposition simulations. The washout ratio 25 offers a method of dealing with this erro r: CMAQ output can be used to calculate w i r values and 26Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-18 August 2008 precipitation measurements (e.g. from NADP) can be used to obtain p 0 values for input to 1 Equation 24 to obtain measurement-corre cted wet deposition velocities. 2
8.3.2.2.2 Aggregation Issues 3
The deposition velocities and washout ratio are dynamic quantities that depend on 4 numerous conditions that are highl y variable with both space and time. However, an ecologically 5 relevant standard is likely to be based on an averaging time greater than hourly, which might 6 require temporal aggregation of deposition veloci ties. Use of deposition velocities that reflect a 7 spatial scale consistent with this review might al so require some level of spatial aggregation of 8 deposition velocities. The v dry

, v wet , and w i r parameters are all expres sible as ratios of other 9 quantities, which allow for multiple possible met hods of calculating aggregate values, including 10 the following: 11
1. Average ratio: Averaging hourly values of the deposition velocities and the washout 12 ratio is equivalent to find the averag e ratios of their component quantities. 13 2. Weighted Average Ratio: Same as above but weight the hour-speci fic ratios by either 14 concentration or deposition dur ing the averaging process. 15 3. Ratio of averages: Instead of worki ng with hourly deposition velocities and w i r 16 values, concentrations and depositions can be aggregated (concentrations are 17 averaged, depositions and precipitation are summed) over space and/or time and the 18 ratios of the resulting aggr egate values can be comput ed. This method removes the 19 outlier effect of unrealistically large ra tios resulting from small denominators (this 20 seems most likely to be an issue in calculations of w i r when precipitation is very small 21 but non-zero). 22 4. Conservatism: Single hourly values from the pool of available data can be chosen that 23 provide a more conservative calculation of depositions, such as predicting the largest 24 possible depositions from a given concentr ation (i.e., choose the annual and spatial 25 maximum values of v i dry and/or w i r for each species and region). 26
Temporal aggregation is known to produce subs tantial biases in dry depositions when 27

correlation exists between v dry and C Amb i (Clarke et al., 1997), and it is presumed that this bias 28 needs consideration when performing spatial ag gregation as well. Some exploration of the 29 conservative method of finding v dry values has been performed, and it was found that this 30 produced unrealistically high a nnually representative values. For purposes of the example 31 calculations that will be presen ted in Section 8.4, the Ratio of Averages method will be 32 employed to obtain all annual depos ition velocities, but the best method of obtaining such values 33 remains an open research question. 34Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-19 August 2008 8.3.2.3 Results and Findings 1 Potential applications of the transforma tion function to the standard setting and 2 implementation processes will be presented in Secti on 8.4. In this section, the variability of the 3 deposition velocities resulting from CMAQ output is explored to 1) highlight the complexity of 4 the process of converting between ambient concentrations and deposit ions and 2) illustrate some 5 possible visualization techniques 6 Fields of dry deposition velocities of SO 2 output by MCIP are shown in Figure 8.3-2

. 7 The figure shows maps of mean and standard de viation of hourly dry deposition velocities over 8 24 hours during four seasonally distinct days ac ross the eastern U.S. output by MCIP. The plots 9
show that v dry values vary by at least two orders of magnitude over the spatial domain, and that 10 there appears to be some spatial co rrelation between the magnitude of v dry values (mean plots) 11 and the hourly variability (standa rd deviation plots). These maps highlight the difficulty of 12 choosing a single nationally representative v dry value. 13 Therefore, v dry values could also be studied at smaller spatial scales. Figures 8.3-3 and 14
8.3-4 show distributions of v dry values from a Case Study Area (Pamlico Sound, NC) of the same 15 four days used for Figure 8.3-2. v dry values in Figure 8.3-3 show an increasing trend in the 16 January 1 plot, a decreasing trend on April 1, and relatively stable trends on July 1 and October 17 1, but with much greater spatia l variability in the July 1 plot . Figure 8.3-4 shows maps of the 18 diurnal distributions of the same data, and reve als that a substantial portion of the spatial 19 variability might be explained by differences in the dominant land use of the grid cells (see 20 Appendix A for the full list of land use categories in the model), especially when distinguishing 21 between water and land surfaces. This suggests that v dry values specific to at least the land use 22 type, if not the grid cell, should be consid ered when using the tr ansformation function. 23Chapter 8 – Considerations in the Structure of the NOx/SOxSecondary Standard(s)DRAFT 8-20 August 200812Figure 8.3-2:Mean and standard deviations ofSO2vdryvalues across the CMAQ eastern United States modeling domain for the3first day of four seasonally distinct months. Note that the high value of the legend represents the maximum of the scale, not of the4data.5Chapter 8 – Considerations in the Structure of the NOx/SOxSecondary Standard(s)DRAFT 8-21 August 200812Figure 8.3-3:Hourly distributions ofSO2vdryvalues for one HUC (#3020105: Pamlico Sound, NC) over the first day of four3seasonally distinct months.4Chapter 8 – Considerations in the Structure of the NOx/SOxSecondary Standard(s)DRAFT 8-22 August 200812Figure 8.3-4:Distributions ofSO2vdryin each gridcell of one HUC (#3020105: Pamlico Sound, NC) over the first day of four3seasonally distinct months. Colors indicate the dominant land use category (DLUSE) of the gridcell; seeAppendix Afor a4description of each categorical value.5Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-23 August 2008 1
Figure 8.3-5 shows domain-wide SO 2 washout ratios values for the same four seasonally 2 distinct days as above, calculat ed using the Average Ratio and the Ratio of Averages methods. 3 These plots illustrate that the potential for unrealis tically extreme values is far greater using the 4 Average Ratio method. Exploration of the numbers in these calculations showed that, despite 5 excluding times when p 0 was 0, very low ambient concentration values could also cause extreme 6 values such as those shown in Figure 8.3-5. These results thus suggest that the Ratio of Averages 7 might be a better choice for calculations of w r i rather than the Average ratio method to help 8 mitigate the use of unrealistically large outlier values. 9

8.3.2.4 Uncertainties 10 The most prominent uncertainty in the process of converting between concentrations and 11 depositions is the known lack of accuracy of the deposition velocities that are input to Equation 12
23. The CASTNET and NADP networks measure c oncurrent weekly values of wet deposition 13 and ambient concentrations of select species; th ese data might be used to estimate the bias 14 present in calculations of v Wet i . Such data explorations are al so expected to aid in understanding 15 the stability of w r i values, which might be low due to the dependence of cloud chemistry 16 dynamics on meteorology of the upper atmosphere . Estimates of deposition velocities might be 17 improved over CMAQ predictions by packaging th e relevant portions of the MCIP and CMAQ 18
source code into a stand-alone tool that generates values using local meteorological 19

measurements (as opposed to the simulated values that are used during model runs). Further 20 discussions of uncertainty will be included in future drafts of this document. 21Chapter 8 – Considerations in the Structure of the NOx/SOxSecondary Standard(s)DRAFT 8-24 August 200812Figure 8.3-5:Daily average washout ratios ofSO2across the CMAQ eastern United Statesmodeling domain for the first day of3four seasonally distinct months (calculated using two different aggregation methods).4Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-25 August 2008 8.3.3 Linkages between Deposition and Ecological Indicators 1 The equation defining ANC forms the basis of the linkage between deposition and 2 surface water acidic condition and the modeling approach used. In the discussion below, the 3 functions are framed in terms of a maximum depositional load given a specified ANC limit, for 4 ease of discussion. However the maximum depo sitional load is not the deposition metric 5 employed in Fig 8.1. The deposition metric is th e deposition associated wi th a certain measured 6 or assumed level of atmospheric concentra tion of NOx and SOx, and under Fig 8.1 the 7 deposition metric is used to calculate an ANC leve l associated with that atmospheric level. The 8 next draft will also frame the functions in terms of calculating an ANC level from a specific 9 deposition metric, whether it is ab ove or below the depositional load. 10 Given an ANC concentration [ANC level ]), the corresponding depositional load (DL) is 11 simply the input flux of acid anions from atmospheric deposition, which gives the [ANC level ] 12 when subtracted from the sustainable acidity or charge balance, defined as the uptake (N upt ), 13 immobilization (N imm ), and denitrification (N den ) of NO x deposition, the in-l ake retention of NO x 14 deposition (N re ) and SO x deposition (S ret ), and the acid anion leaching from the catchment as a 15 function of the pre-industria l flux of base cations [BC] 0 * from weathering, a described in 16 Equation 5: 17
18 DL(N NOx ) + DL(S SOx ) ={ f

N upt + (1 - r)(N imm + N den ) + r(N ret + S ret )} + ( [BC] 0 * - [ANC level ])Q (5) 19
20 where f and r are dimensionless parameters that define the fraction of forest cover in the 21 catchment and the lake/catchment ratio, respec tively. The right hand side of equation (5) is 22 equivalent to the g(·) function di scussed in Section 8.3.1. To c onvert into fluxes, concentrations 23 ([x] = X/Q) are simply multiplied by runoff (Q ) (in m/yr) from the site. The charge balance 24 equates the deposition inputs of acid anions with the sum of pr ocesses which control their long 25 term storage, removal and leaching exports. Seve ral major assumptions are made: 1) steady-state 26 conditions, 2) the effect of nutri ent cycling between plants and soil is ignored 3) there are no 27 significant N inputs from sources other than atmospheric deposition, 4) ammonium leaching is 28 negligible because any inputs are either taken up by biota, adsorbed onto soils or nitrate 29 compounds, and 5) long-term sinks of sulfate in the catchment soils are negligible. 30Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-26 August 2008 Using equation 5 to derive the terms for intern al catchment processes, the charge balance 1 provides those combinations of N NOx and S SOx deposition that defines th e DL function given by: 2
3

DL max (S SOx ) = (( [BC] 0 * - [ANC level ])Q)/ (1- p s ) (6) 4

DL min (S SOx ) = 0 (7) 5

DL max (N NOx ) = fN upt + (1-r)( N imm + N den ) + [( [BC] 0 * - [ANC level ])Q]/ (1- p n ) (8) 6

DL min (N NOx ) = fN upt + (1-r)( N imm + N den ) (9) 7

where p s and p n define the fraction of in-lake retention of N ret and S ret, which is similarly 8 proportional to the net input of N NOx and S SOx to the lake. It is not possible to define a single 9 value of maximal loading of total acidity, sin ce the acid anions sulfate and nitrate behave 10 differently in the way they are transported wi th hydrogen ions; one unit of deposition of S will 11
not have the same net effect on surface water ANC as an equivalent unit of N deposition. 12 Equations 2 and 4 define the deposition loads for N NOx and S SOx individually. DLmax(S SOx ) 13 defines the loading for sulfur when total N NOx deposition is less then DLmin(N NOx ). 14 DLmin(N NOx ) defines the deposition of N NOx at which terrestrial catc hment processes effectively 15
remove N NOx , so that deposition loads lower then DLmin(N NOx ) result in no leaching of nitrate. 16 DLmax(N NOx ) defines the load for total N NOx deposition when S deposition is zero. In reality, 17
neither N NOx and S SOx deposition will ever be zero, so the DL for the deposition of one is fixed 18 by the deposition of the other, acco rding to the line defining in Figure 8.3-1 . It is our intention to 19 construct a similar line for N deposition originating from NO x only for the second draft of this 20
risk assessment. 21Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-27 August 2008 N NOx Deposition DL max (N NOx )SSOxDeposition

DL min (N NOx )DLmax(SSOx)

N NOx Deposition DL max (N NOx )SSOxDeposition

DL min (N NOx )DLmax(SSOx)

1 Figure 8.3-1 . The deposition load function defined by the model. 2

The thick lines indicate all possible pairs of DL of N NOx and S SOx acidity. Note that in the 3
above formulation individual DL of N NOx and S SOx are not specified; each pair of depositions 4 (N NOx and S SOx ) fulfilling equations 2-5. 5 Given a [ANC level ] for a desired level of protection, the modeling has to provide the 6
calculations of the N NOx and S SOx sinks (N upt , N imm, N den , N ret , S ret ), landscape characteristics (f 7 and r) and critical leachi ng of acid anions (( [BC] 0 * - [ANC level ])Q), which are described below. 8
9 Given a maximum deposition load, ANC concentr ation at a site can be provided by the 10
difference between future base cation and acid anion concentrations, which is equal to the 11 predicted, steady-state sulfate (S leach ) and nitrate (N leach ) leaching fluxes: 12
[ANC] = [BC] 0 * - (S leach + N leach )/Q (10) 13

where S leach and N leach are: 14

S leach = (1 – p s )S dep (11) 15

N leach = (1- p n )(N dep – f N upt – (1 – r)(N imm + N den )) (12) 16

17

18Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-28 August 2008 8.3.3.1 Parameters 1 Net growth uptake of N NOx in forest (N upt ) 2 At steady-state or over long timescal es, the only permanent sink for N (from any source) 3 in the terrestrial system occurs when there is removal of biomass from the catchment. Short term, 4 seasonal cycling of N does not affect the mean annual budget over these timescales. The main 5 removal of N from the catchment is through the harvesting of the forest or grazing. There are 6 various ways for estimating this N loss from the catchment, which take into account the N 7 content of various tree species and the time periods of harvesti ng cycle. For forests without 8 harvesting, like the Adirondack region of New York, N upt is zero. 9
10 Immobilization (N imm ) and denitrification (N den ) of N NOx in catchment soils 11 For long-term immobilization of N is forest so ils, a constant value of 2 kg N / ha/ yr is 12 used. This value represents the lower end of the range suggested from studies (Downing et al 13 1993). Thus, the amount of N removed for a given N NOx depositional load is simply 14 propositional to its total N tot load (N NOx /N tot ) 15

It is assumed that the denitrificat ion is related to the net input of N NOx loading into the soil 16 system after accounting for the removal of N NOx by immobilization and uptake as follows: 17

N den = f de(N nox – N imm – N upt ) (13) 18

where f de is the denitrification faction (value between 0 and 1), ascribed to the catchment soil. 19 Denitrification is an anaerobic process and therefore a function of soil moisture content, with 20 higher rates in peat soils. A t ypical upland soils denitrify N at a rate of 10% of N input while 21 peat soils work at a rate of 70 to 80 %. Ther efore, the average denitr ification fraction for a 22 catchment is approximated by the following linear relationship: 23
f de = 0.1 + 0.7 f

peat (14) 24

where f peat is the proportion of peat soils within the catchment. 25
26
In-lake retention of N NOx and S SOx 27
The in-lake retention of acid anions is assumed to be a linear function of ne t input of acidity. It 28 is sequentially the final sink of acidity encounter ed by inputs that move through the catchment. 29Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-29 August 2008 The in-lake retention fractions for N NOx ( p n ) and S SOx ( p s ) are calculated from a kinetic equation 1 accounting for water retention time (Kelly et al 1987; Posch et al 1997). 2
3

p n = (s n / s n + Q/r) and Ps = (s s / s s + Q/r) (15) 4
5 where Q is the runoff and s n and s s are the net mass transfer coefficient for N NOx and S SOx . Due to 6 a lack of data, the net mass tran sfer coefficient for sulfate (s s ) was taken from a retention model 7
calibration to mass balance data of 11 lakes located in Nort h America and northern Europe 8 (Baker and Brezonik 1988). The mean value of s s for these lakes, 0.5 m/yr, was used for the 9 maximum depositional load calculations. The value for s n , 5 m/yr, was obtained from a similar 10 study of 12, mostly Canadian, lakes (Dillon and Molot 1990). 11

Pre-Industrial Base Cation Concentration 12 The pre-industrial concentration of base cations [BC] 0 * effectively set the long term 13 capacity of the catchment to neutralize acidic depo sition, because it repres ents the only source of 14 base cation input that is sustainable over th e long-term. Input of cations from weathering is 15 assumed to be a relatively constant pr ocess driven largely by the reaction of CO 2 with primary 16
minerals in the soils and bedrock. Base cati ons are removed by leaching from the soil solution 17 through surface water runoff. At a steady-state, the l eaching rate of base cation occurs at lesser or 18 greater rates than the weathering supply. Howeve r, base cation leaching is not at steady-state 19 today because anthropogenic acid deposition actu ally increases the leaching of base cations 20 through ion-exchange within catchment soils. Soils contain a store of adso rbed base cations, as 21 measured as base saturation, which are derive d from weathering, but have accumulated in the 22 soil over millennia, until eventua lly a steady-state is achieved, whereby the supply of base 23 cations from weathering was in approximate equi librium with the removal of base cations by 24 rainwater, itself in equilibrium with the atmosphere. For this reason, [BC] 0 * cannot be derived 25 from measured data in runoff, but derived from a empirical relationships (i .e., pre-industrial base 26 cation concentration). 27 The pre-industrial base cation concen tration is the sum of weathering ([BC *
w ]) supply 28 plus base cation deposition ([BC * dep ]), if it is assumed that ba se cation deposition has not 29Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-30 August 2008 significantly changed since pre-i ndustrial times, minus long-term average uptake of base cations 1 in the biomass (i.e., the annual average rem oval of base cations due to harvesting): 2
[BC] 0 * = [BC w ] + [BC * dep ] – [BC u ]. (16) 3
The first step is to quantify the proportion of measured base cation leaching which is 4
derived from transient i on-exchange processes (BC ex ) and is proportional to the load of acid 5 anions. This proposition is represented by the term ‘F’, calculated according to the methodology 6 of Brakke and colleagues (1990): 7
F = sin ( π /2
. [BC * ] t /[S] (17) 8
where [BC * ] t is the measured non-marine base cation c oncentration and S is a constant which 9 varies regionally according to geology, but from empirical studies is taken as 400 μ eq/L 10 (Harriman and Christie, 1995). This constant determines the m easured non-marine base cation 11 concentration which represents a catchment lik ely to be unaffected by acid deposition; [BC * ] t = 12 S, F =1 and base cation leaching is increased by exactly the value of the acid anion load, 13 resulting in no change in the ANC of runoff. For values of [BC * ] t greater then S, F is 1. 14 This F value is then used to calculate the pre-industrial base cation concentration 15 according to the following equation: 16
[BC * ] 0 * = [BC * ] t – F([AA] t * - [AA] 0 * ) (18) 17
where [AA] 0 * is the pre-acidification concentration of non-marine acid ani ons from weathering 18 and natural atmospheric sources and the measur ed leaching rate of non-marine base cations 19 [BC * ] t represent the sum of weathering, non-mari ne deposition and leaching sources. We 20 assumed the steady-state concentration of nitrate ([AA] 0 ) was zero ([AA] 0 * = 0). Since [BC * ] 0 * is 21 now known, the maximum depositional load can be defined by Equation 9. 22

Catchment Parameters (r, f, and Q) 23
Lake and catchment areas can be estimated from topographical maps (1:50,000) and from 24 the EMAP lake survey. The annual runoff can be obtained by interpol ation from a national 25 runoff map (Krug et al., 1990) for water years 1951- 80. The fraction of forest in the catchment 26
can be estimated from 2001 Nati onal Land Cover Data (NLCD). 27Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-31 August 2008 

### 8.4 EXAMPLE APPLICATIO NS OF SECONDARY NAAQS 1 

### STRUCTURE 2 This section is intended to provide an illust rative example applying actual data sets for 3 the Adirondacks to the above derivations and linkag e calculations. It is our intent that this 4 section will be developed in the second draft of this document. 5



### 8.5 REFERENCES 6 Baker, L. A., and P. L. Brezonik. 1988. Dynami c model of in-lake alkali nity generation. Water 7
Resources Research 24: 65–74. 8

Brakke, D.F., A. Henriksen, and S.A. Norton. 1990. A variable F-factor to explain changes in 9
base cation concentrations as a function of strong acid deposit ion. Verh. Internat. Verein. 10
Limnol. 24:146–149. 11

Clarke, J., Edgerton, E., Martin, B., 1997. Dry de position calculations for the Clean Air Status 12 and Trends Network. Atmosphe ric Environment 31 (21), 3667–3678. 13

Davis, J. M., Swall, J. L., 2006. An examination of the CMAQ simulations of the wet deposition 14 of ammonium from a Bayesian perspec tive. Atmospheric Environment 40, 4562–4573. 15

Dillon, P. J., and L. A. Molot. 1990. The role of ammonium and nitrate in the acidification of 16
lakes and forested catchment s. Biogeochemistry 11:23–43. 17

Downing, R. J., J.-P. Hettelingh, and P. A. M. de Smet (eds.). 1993. Calculation and mapping of 18 critical loads in Europe, st atus report 1993. Coordination Ce nter for Effects, National 19
Institute of Public Health and Enviro nmental Protection (RIVM), Bilthoven, The 20
Netherlands. 21

Harriman R, Christie AEG. 1995 Estimating criti cal loads for biota: the steady-state water 22 chemistry Henriksen.model. Critical Lo ads of Acid Deposition for United Kingdom 23 Freshwaters:7_8. 24Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-32 August 2008 Kelly, C. A., J. W. M. Rudd, R. H. Hesslin, D. W. Schindler, P. J. Dillon, C. T. Driscoll, S. A. 1 Gherini, and R. H. Heskey. 1987. Prediction of biological acid ne utralization in acid 2 sensitive lakes. Biogeochemistry 3:129

– 140. 3

Krug, W. R., W. A. Gebert, D. J. Graczyk, D. L. Stevens, B. P. Rochelle, and M. R. Church, 4 1990. Map of Mean Annual Runoff for the Nort heastern, Southeastern , and Mid-Atlantic 5 United States, Water Years 1951-1980. U.S. Geological Survey Water- Resources 6 Investigations Report 88- 4904, Madison, Wisconsin. 7

Morris, R. E., McNally, D. E., Tesche, T. W., Tonnesen, G., Boylan, J. W., Brewer, P., 2005. 8 Preliminary evaluation of the community mu ltiscale air quality model for 2002 over the 9
Southeastern United States. Journal of the Air and Waste Management Association 55, 10
1694–1708. 11

Posch, M, J.K.MA, R. Forsius, A. Henriksen, A. Wilander 1997 Exceedance of Critical Loads 12 for Lakes in Finland, Norway, and Swed en: Reduction Requirements for Acidifying 13 Nitrogen and Sulfur Deposition. Environmen tal Management Vol. 21, No. 2, pp. 291– 14
304. 15

Seinfeld, J., Pandis, S., 1998. Atmospheric Chemis try and Physics. John Wiley and Sons, Inc., 16
New York. 17

U.S. EPA. 2008. Integrated Science Assessment for Oxides of Nitrogen and Sulfur – 18 Environmental Criteria (Second External Review Draft) . Prepared by U.S. Environmental 19 Protection Agency, National Center for Envi ronmental Assessment-R TP Division, Office 20 of Research and Development, Research Triangle Park, NC. EPA/600/R-08-082. August. 21

U.S. EPA, 1999. Science Algorithms of the EP A Models-3 Community Multiscale Air Quality 22 (CMAQ) Modeling System. Tech. Rep. EPA/600/R-99/030, U.S. Environmental 23 Protection Agency, Washington DC. 24
25 6 7

28Chapter 8 – Considerations in the Structure of the NO x /SO x Secondary Standard(s)

DRAFT 8-33 August 2008 

### APPENDIX A 1 Land Use Categories in MCIP (from USGS). 2
1. Urban Land 3 2. Dryland Cropland and Pasture 4 3. Irrigated Cropland and Pasture 5 4. Mixed Dryland and Irrigated Cropland 6 5. Cropland/Grassland Mosaic 7 6. Cropland/Woodland Mosaic 8 7. Grassland 9 8. Shrubland 10 9. Mixed Shrubland/Grassland 11 10. Savannah 12 11. Deciduous Broadleaf Forest 13 12. Deciduous Needleleaf Forest 14 13. Evergreen Broadleaf Forest 15 14. Evergreen Needleleaf Forest 16 15. Mixed Forest 17 16. Water 18 17. Herbaceous Wetland 19 18. Wooded Wetland 20 19. Barren or Sparsely Vegetated 21 20. Herbaceous Tundra 22 21. Wooded Tundra 23 22. Mixed Tundra 24 23. Bare Ground Tundra 25 24. Snow or Ice 26 27United StatesEnvironmental ProtectionAgencyOffice of Air Quality Planning and StandardsHealth and Environmental Impacts DivisionResearch Triangle Park, NCPublication No. EPA-452/P-08-005bAugust 2008