&EPA
United State
Environmental PwtocSon
Agency
Policy Assessment for the Review of
the Secondary National Ambient Air
Quality Standards for Oxides of
Nitrogen and Oxides of Sulfur

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                                             EPA-452/R-ll-005a
                                                 February 2011
  Policy Assessment for the Review of the Secondary
National Ambient Air Quality Standards for Oxides of
             Nitrogen and Oxides of Sulfur
                 U.S. Environmental Protection Agency
               Office of Air Quality Planning and Standards
               Health and Environmental Impacts Division
                Research Triangle Park, North Carolina

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                                    DISCLAIMER

       This document has been reviewed by the Office of Air Quality Planning and Standards,
U.S. Environmental Protection Agency, and approved for publication.  This final document has
been prepared by staff from the Office of Air Quality Planning and Standards, U.S.
Environmental Protection Agency.  Any opinions, findings, conclusions, or recommendations are
those of the authors and do not necessarily reflect the views of the U.S. Environmental
Protection Agency. Mention of trade names or commercial products is not intended to constitute
endorsement or recommendation for use. Any questions or comments concerning this document
should be addressed to Richard Scheffe, U.S. Environmental Protection Agency, Office of Air
Quality Planning and Standards, C304-02, Research Triangle Park, North Carolina 27711 (email:
scheffe.richard@epa.gov).

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                               ACKNOWLEDGMENTS

       This Policy Assessment is the product of the Environmental Protection Agency's (EPA)
Office of Air Quality Planning and Standards (OAQPS), with important contributions from the
Office of Atmospheric Programs (OAP) and the Office of Research and Development's (ORD)
National Center for Environmental Assessment (NCEA).  Richard Scheffe led the team that was
responsible for the development of this document, which built upon earlier drafts of the
document that had been developed under the leadership of Bryan Hubbell.  The principal authors
of chapter 2 on emissions, air and water quality are Richard Scheffe, Jason Lynch and Norm
Possiel. The principal authors of chapter 3 on deposition-related effects are Tara Greaver
(NCEA) and Meredith Lassiter (NCEA). The principal authors of chapter 4 on adversity to
public welfare are Christine Davis, David Evans (EPA,  OP) and Brian Heninger (EPA, OP). The
principal author of chapter 5 on co-protection is Randy Waite. The principal authors of chapter 6
on adequacy of the standards are Ginger Tennant and Bryan Hubbell. The principal authors on
chapter 7, elements of the standard, are Richard Scheffe and Karen Martin. Appendix A,
analysis of critical loads, was developed under contract to RTI.  Appendix B,  critical load
modeling, was authored by Jason Lynch (OAP) and Tara Greaver (NCEA). Appendix C,
ecoregion atlas, was authored by Travis Smith. Appendix D,  alternative standards, and
Appendix E, derivation of nitrate as an indicator for NOy, were authored by Richard Scheffe.
Appendix F on uncertainty was authored by Richard Scheffe,  Adam Reff, Travis Smith, and
Norm Possiel.  Appendix G on cumulative uncertainty analysis was authored  by Travis Smith
and Lyle Burgoon (ORD-NHEERL).  Critical review and editing of the document were provided
by Richard Scheffe, Travis Smith, Ginger Tennant, Randy Waite, and Karen Martin. Valuable
comments were also provided by John Hannon, Lea Anderson, and Steve Silverman from EPA's
Office of General Counsel.

       The principal analysts responsible for most of the computer programming, data base
management and calculations were Adam Reff, Jason Lynch,  Tara Greaver, Bryan Hubbell, and
Travis Smith  Additional technical assistance was provided by Robin Dennis, Fred Dimmick,
Kristen Foley, Gary Lear, Russell Long, Rob Finder, Joe Sickles, Joe Tikvart, and John Walker .
Especially noted are the conceptual contributions of Jason Lynch that led to the development of a
national data base of critical loads that provided a major infrastructure component for this
assessment. We also acknowledge the significant contributions of Anne Rea, who led the
development of the Risk and Exposure Assessement, and Tara Greaver, who led the
development of the Integrated Science Assessment, upon which this document is based.
Jacinthe Racine of the Canadian Meteorological Centre provided AURAMS air quality and
deposition data.

       Earlier drafts of this document were reviewed by the Clean Air Scientific Advisory
Committee (CASAC). This document has been informed by the expert advice and comments
received from CASAC, as well as by public comments.

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Policy Assessment for the Review of the Secondary National
Ambient Air Quality Standards for Oxides of Nitrogen and Sulfur
                             EXECUTIVE SUMMARY
Introduction
This Policy Assessment has been prepared by staff in the Environmental Protection Agency's
(EPA) Office of Air Quality Planning and Standards (OAQPS) in conjunction with the Agency's
ongoing joint review of the secondary (welfare-based) national ambient air quality standards
(NAAQS) for oxides of nitrogen and sulfur. This Policy Assessment evaluates the policy
implications of the key scientific information contained in the Integrated Science Assessment
(ISA) for Oxides of Nitrogen and Sulfur-Ecological Criteria, prepared by EPA's National Center
for Environmental Assessment (NCEA), results from the analyses contained in the Risk and
Exposure Assessment (REA) for Review of the Secondary National Ambient Air Quality
Standards for Oxides of Nitrogen and Oxides of Sulfur, as well as additional staff analyses
conducted for and presented in this document. It presents staff conclusions regarding the
adequacy of the current NO2 and SO2 secondary standards as well as alternative standards for
consideration in this review.
This Policy Assessment is intended to help "bridge the gap" between the relevant scientific and
technical information and the judgments required of the EPA Administrator in determining
whether, and if so, how, it is appropriate to revise the secondary NAAQS for oxides of nitrogen
and sulfur. This Policy Assessment considers the available scientific evidence and quantitative
risk-based analyses, together with related limitations and uncertainties, and focuses on the basic
elements of air quality  standards: indicator, averaging time, form, and level. These elements,
which serve to define each standard, must be considered collectively in evaluating the public
welfare protection afforded by the standards.

Scope
In conducting this periodic review of the secondary NAAQS for oxides of nitrogen and sulfur,
EPA has decided to jointly assess the scientific information, associated risks, and standards
because oxides of nitrogen and sulfur in the ambient air, and their associated transformation
products, such as  deposited nitrogen and sulfur, are linked from an atmospheric chemistry
perspective, as well as jointly contributing to ecological effects.
For this Policy Assessment, we have chosen to focus much of our attention on effects in sensitive
aquatic ecosystems caused by acidifying deposition of nitrogen and sulfur, which is a
transformation product of oxides of nitrogen and sulfur in the ambient air. We have a high
degree of confidence in the linkages between atmospheric oxides of nitrogen and sulfur,
associated deposition of nitrogen and sulfur, and deposition-related aquatic acidification effects.
Our objective in this Policy Assessment is to develop a framework for a multi-pollutant,
multimedia standard that is ecologically relevant and reflects the combined impacts of these two
pollutants as they deposit to sensitive aquatic ecosystems.
In so doing, we recognize that a standard developed specifically to address aquatic acidification
would not likely provide targeted protection against other deposition-related ecological effects,
including effects related to terrestrial acidification and nutrient enrichment effects in sensitive
terrestrial and aquatic ecosystems. Nonetheless, it is likely that some additional protection from
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Policy Assessment for the Review of the Secondary National
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these other effects would result from reductions in atmospheric oxides of nitrogen and sulfur that
would likely occur in response to an ecologically relevant aquatic acidification standard.
 In this Policy Assessment we use the term total reactive oxidized nitrogen, NOy, as used by the
scientific community, to represent the complete set of oxidized nitrogen compounds.  The major
gaseous and paniculate constituents of NOy include nitric oxide (NO), nitrogen dioxide (NO2),
nitric acid (HNOs), peroxyacetyl nitrate (PAN), nitrous acid (HONO), organic nitrates, and
particulate nitrate (NOs). In contrast, the term NOx more narrowly refers to the sum of NO2 and
NO.  Total oxides of sulfur include both gaseous substances [e.g., sulfur dioxide (802), sulfur
monoxide (SO), sulfur trioxide (SOs), thiosulfate (8203), and heptoxide (820?)], as well as
particulate species,  such as ammonium sulfate [(NH4)2SO4]. Throughout this document, we
focus more narrowly on SOx, defined as the sum of SO2 and particulate sulfate (804), which
represent virtually all of the oxidized sulfur mass in the atmosphere.

Deposition-related Ecological Effects Associated with Oxides of Nitrogen and Sulfur
Deposition-related ecological effects are broadly categorized into those related to acidification
and nutrient enrichment.  Acidification occurs in both aquatic and terrestrial ecosystems, with
most aquatic effects occurring in freshwater lakes and streams. Nutrient enrichment also occurs
in both aquatic and terrestrial ecosystems; however, the types and prevalence of nutrient
enrichment effects vary between freshwater and estuarine aquatic ecosystems.
In the acidification process, geochemical components of terrestrial and freshwater aquatic
ecosystems are altered in a way that leads to effects on biological organisms. Because oxides of
nitrogen  and sulfur deposited to terrestrial ecosystems often move through the soil and
eventually leach into adjacent water bodies, deposition to terrestrial ecosystems is also a cause of
acidification in aquatic ecosystems.
The scientific  evidence is sufficient to infer a  strong causal relationship between acidifying
deposition and effects on biogeochemical processes and biota in aquatic ecosystems, and
between  acidifying deposition and  changes in biogeochemistry in terrestrial ecosystems.
Acidifying deposition is observed to alter sulfate and nitrate concentrations in surface waters,
acid neutralizing capacity (ANC), inorganic aluminum, and surface water pH. These changes
can result in the loss of acid-sensitive biological species such as salmonids and disrupt food web
dynamics causing alteration to the diet, breeding distribution and reproduction of certain species
of bird, such as goldeneye ducks and loons. Acidification in terrestrial ecosystems has been
shown to cause decreased growth and  increased susceptibility to disease and injury in sensitive
tree  species, including red spruce and sugar maple.
Principal factors governing the sensitivity of terrestrial and aquatic  ecosystems to acidification
from sulfur and nitrogen deposition include geology, plant uptake of nitrogen, soil depth, and
elevation. Geologic formations having low base cation supply generally underlie the watersheds
of acid-sensitive lakes and streams. Other factors that contribute to the sensitivity of soils and
surface waters to acidifying deposition include topography, soil chemistry, land use, and
hydrologic flowpath. Chronic as well as episodic acidification tends to occur primarily at
relatively high elevations in areas that have base-poor bedrock, high relief, and shallow soils.
With regard to aquatic acidification, based on analyses of surface water data from freshwater
ecosystem surveys and monitoring, the most sensitive lakes and streams are contained in New
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England, the Adirondack Mountains, the Appalachian Mountains (northern Appalachian Plateau
and Ridge/Blue Ridge region), the mountainous West, and the Upper Midwest.
ANC is the most widely used indicator of acid sensitivity and has been found in various studies
to be the best single indicator of the biological response and health of aquatic communities in
acid sensitive systems.  Annual or multi-year average ANC is a good overall indicator of
sensitivity, capturing the ability of an ecosystem to withstand chronic acidification as well as
episodic  events such as spring melting that can lower ANC over shorter time spans. Biota are
generally not harmed when annual average ANC levels are > 100 microequivalents per liter
(|ieq/L).  At annual average ANC levels between 100 and 50 |ieq/L, the fitness of sensitive
species (e.g., brook trout, zooplankton) begins to decline. When annual average ANC is <50
ueq/L, negative effects on aquatic biota are observed, including large reductions in diversity of
fish species, and declines in health offish populations, affecting reproductive ability and fitness.
Annual average ANC levels below 0 jieq/L are generally associated with complete loss offish
species and other biota that are sensitive to acidification.  An example of the relationship
between  ANC level and aquatic effects based on lakes in the Adirondacks is illustrated in the
following figure.
                               Severe  Elevated  Moderate
   14
 O 12
   10
  8
 «  6
                     0
                      2
                             Acute
                                                     Low
                                                /
                               i      I
                       -200  -100    0    10O  200   300   400   500
                                       ANC
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Policy Assessment for the Review of the Secondary National
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health of at least a portion of the sugar maple and red spruce growing in the United States may
have been compromised by acidifying total nitrogen and sulfur deposition in recent years. A
commonly used indicator of terrestrial acidification is the base cation-to-aluminum ratio, Bc/Al.
Many locations in sensitive areas of the U.S. have Bc/Al levels below benchmark levels we have
classified as providing low to intermediate levels of protection to tree health. At a Bc/Al ratio of
1.2 (intermediate level of protection), red spruce growth can be reduced by 20 percent. At a
Bc/Al ratio of 0.6 (low level of protection), sugar maple growth can be reduced by 20 percent.
While not defining whether a 20 percent reduction in growth can be considered significant,
existing economic studies suggest that avoiding significant declines in the health of spruce and
sugar maple forests may be worth billions of dollars to residents  of the Eastern U.S.
With regard to nutrient enrichment, the numerous ecosystem types that occur across the U.S.
have a broad range of sensitivity to nitrogen deposition. Organisms in their natural environment
are commonly adapted to a specific regime of nutrient availability. Change in the availability of
one important nutrient, such as nitrogen, may result in imbalances in ecosystems, with effects on
ecosystem processes, structure and function.  In certain nitrogen-limited ecosystems, including
many ecosystems managed for commercial production, nitrogen  deposition can result in
beneficial increases in productivity. Nutrient enrichment effects from deposition of oxides of
nitrogen are difficult to disentangle from overall effects of nitrogen enrichment. This is caused
by two factors:  the inputs of reduced nitrogen from deposition and, in estuarine ecosystems,  a
large fraction of nitrogen inputs from non-atmospheric sources.

Adequacy of the Existing NOi and SOi Standards
Current NC>2 and SC>2 secondary standards are designed to protect against direct exposure of
vegetation to ambient concentrations of oxides of nitrogen and sulfur.  All areas of the U.S.
currently meet the existing NC>2 and 862 secondary standards. The NC>2 secondary standard is
0.053 parts per million (ppm), annual arithmetic average, calculated as the arithmetic mean of
the 1-hour NO2 concentrations.   The SC>2 secondary standard is a 3-hour average of 0.5 ppm, not
to be exceeded more than once per year. Based on currently available information, staff
concludes that the current secondary standards serve to protect vegetation from direct damage
associated with exposures to gaseous SC>2 and NC>2 and thus consideration should be given to
retaining the current standards for that purpose.
With regard to aquatic acidification, recent data indicate that in the Adirondacks and
Shenandoah areas, rates of acidifying deposition of oxides of nitrogen and sulfur are still well
above pre-acidification (1860) conditions. Forty-four percent of Adirondack lakes evaluated
exceed the critical load for an ANC of 50 ueq/L, and in these lakes recreationally important fish
species such as trout are missing due to acidification. In the Shenandoah area,  85 percent of
streams evaluated exceed the  critical load for an ANC of 50 ueq/L, resulting in losses in fitness
in species such as the Blacknose Dace.
With regard to terrestrial acidification, the REA evaluated a small number of sensitive areas as
case studies. In the sugar maple case study area (Kane Experimental Forest, Pennsylvania),
recent (2002) deposition levels are associated with a Bc/Al ratio below 1.2,  indicating the
potential for a greater than 20 percent reduction in growth.  In the red spruce case study area
(Hubbard Brook Experimental Forest, New Hampshire), recent deposition levels are associated
with a Bc/Al ratio slightly above 1.2, indicating slightly less potential for significant reductions
in growth. When the methodology was extended to a 27-state region, the calculated Bc/Al ratio
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Policy Assessment for the Review of the Secondary National
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fell below 1.2 in 12% of the sugar maple plots and 5% of the red spruce plots; however, results
from individual states ranged from 0 to 67% of the plots for sugar maple and 0 to 100% of the
plots for red spruce fell below the Bc/Al ratio of 1.2.
Available ecological indicators for estuarine nutrient enrichment are not sufficiently sensitive to
changes in atmospheric nitrogen oxides to be of use in assessing the adequacy of the current NC>2
secondary standard.  Atmospheric nitrogen oxides can be an important contributor of nitrogen to
estuarine nutrient enrichment, but additional analysis would be required to develop an
appropriate indicator for assessing levels of protection from nutrient enrichment effects in
estuaries related to deposition of nitrogen oxides.
Nitrogen deposition can alter species composition and cause eutrophication in freshwater
systems.  In the Rocky Mountains, for example, deposition loads of 1.5 to 2  kg/ha/yr, which are
within the range associated with ambient nitrogen oxide levels meeting the current standard, are
known to cause changes in species composition in diatom communities indicating impaired
water quality.
With regard to terrestrial nutrient enrichment, most terrestrial ecosystems in the U.S. are
nitrogen-limited, and therefore they are sensitive to perturbation caused by nitrogen additions.
Under recent conditions, nearly all of the known sensitive mixed conifer forest ecosystems
receive total nitrogen deposition levels above 3.1 N kg/ha/yr, which is the ecological benchmark
for changes in lichen species. Lichens are sentinels for broader ecosystem change in terrestrial
systems. Some portions of the Sierra Nevadas receive total nitrogen deposition levels above 5.2
N kg/ha/yr, which is the ecological benchmark for shifts in the dominant species of lichen from
acidophytic to tolerant species.  In addition, in Coastal Sage Scrub ecosystems in California,
nitrogen deposition exceeds the 3.3 N kg/ha/yr benchmark above which nitrogen is no longer a
limiting nutrient, leading to potential alterations in ecosystem composition.
Based on the above considerations, staff concludes that currently available scientific evidence
and assessments clearly call into question the adequacy of the current standards with regard to
deposition-related effects on sensitive aquatic and terrestrial ecosystems, including acidification
and nutrient enrichment.  Further, staff recognizes that the elements of the current standards —
indicator, averaging time, level and form - are not ecologically relevant,  and are thus not
appropriate for standards designed to provide such protection. Thus, staff concludes that
consideration should be given to establishing a new ecologically relevant multi-pollutant,
multimedia standard to provide appropriate protection from deposition-related ecological effects
of oxides of nitrogen and sulfur on sensitive ecosystems, with a focus on protecting against
adverse effects associated with acidifying deposition in sensitive aquatic ecosystems.

Design of an Ecologically Relevant Standard for Aquatic Acidification
The graphic below depicts the framework within which we are considering the structure of an
ecologically relevant secondary standard for aquatic acidification. This conceptual diagram
illustrates how an ecological indicator is linked to concentrations of ambient air indicators of
oxides of nitrogen and sulfur through deposition.
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Policy Assessment for the Review of the Secondary National
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                        Linking atmospheric
                        oxides of S and N deposition to
                        ecological indicator

  Ecological effects and                                  Linking deposition to "allowable"
  ecological indicator                                    concentrations of NOy and SOx
  (ANC)


This Policy Assessment is organized around this conceptual framework.  It presents our current
understanding of the ecological and atmospheric factors that modify the impacts of deposited
oxides of nitrogen and sulfur on sensitive ecosystems. Applying this framework has resulted in
the development of a new ecologically relevant standard that incorporates multi-pollutant and
multimedia attributes in linking ambient air indicators of oxides of nitrogen and sulfur to an
ecological indicator through atmospheric deposition.  There are three main  components of the
conceptual design of the standard:  (1) linkage between ecological indicators and ecological
effects, (2) linkage between an ecological indicator and atmospheric deposition, and (3) linkage
between  deposition and ambient air indicators.
In this Policy Assessment, the focus is on developing a standard that protects against ecological
effects associated with acidifying deposition of oxides of nitrogen and sulfur in aquatic
ecosystems, recognizing that both oxides of nitrogen and sulfur are major contributors to aquatic
acidification and that acidification of aquatic ecosystems is best characterized and understood in
terms of the combined rather than individual effects of oxides of nitrogen and sulfur.  In addition,
there is a well developed body of scientific evidence linking the deposition of ambient oxides of
nitrogen  and sulfur to acidification in sensitive aquatic ecosystems. While we conclude that the
available information and assessments are only sufficient to support the development of a
national  standard specifically to address aquatic acidification at this time, we  recognize that this
general conceptual framework could likely be applied to a broader set of deposition-related
effects in the future.
In focusing on the effects of acidifying deposition on aquatic ecosystems, with respect to linking
ecological indicators to adverse effects offish mortality and decreased species diversity, staff
concludes that ANC  is the most appropriate ecological indicator to consider.  ANC is the most
widely used chemical indicator of acid sensitivity  in aquatic ecosystems and has been found
through numerous studies to be the best single indicator of the biological response and health of
aquatic communities in acid sensitive ecosystems.  Furthermore, ANC can be directly linked to
both underlying water chemistry, e.g. pH and aluminum, and to biological impairment,
specifically fish mortality and the number  offish species in  a water body.
With respect to linking atmospheric deposition to  the ecological indicator,  staff concludes that
steady state ecosystem acidification modeling that calculates critical loads is the appropriate
methodology to link  atmospheric deposition with ANC.  A critical load for acidity is the amount
of acidifying deposition beyond which a water body cannot  achieve and sustain a target ANC
level. Critical loads reflect the relative sensitivity to acidification of a water body within a
distribution of water bodies.
With respect to linking deposition to ambient air concentrations, staff has developed the concept
of transference ratios, which are the ratio of deposition to ambient air concentration, as an
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Policy Assessment for the Review of the Secondary National
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appropriate approach to use in linking deposition to ambient air concentrations.  Representative
transference ratios that are averaged annually and over a specified spatial area have been
developed for oxides of nitrogen and oxides of sulfur based on simulations using EPA's
Community Multiscale Air Quality (CMAQ) model.

Staff Consideration of Alternative Standards for Aquatic Acidification
In applying the framework that reflects these three fundamental linkages, staff has developed an
ecologically relevant standard for aquatic acidification in terms of the basic elements that
together define a NAAQS:  ambient air indicator, form, averaging time, and level.
With regard to ambient air indicators, staff concludes that consideration should be given to using
total reactive oxidized nitrogen, NOy, as the ambient air indicator for oxides of nitrogen and the
sum of gaseous sulfur dioxide (862) and particulate sulfate (SO/j), referred to in this assessment
as SOX, as the ambient air indicator for oxides of sulfur,
With regard to the form of such a multi-pollutant, deposition-related standard, staff concludes
that consideration should be given to an ecologically relevant form that characterizes the
relationships between the ambient  air indicators for oxides of nitrogen and sulfur, the related
deposition of nitrogen and sulfur, and the associated aquatic acidification effects in terms of the
ecological indicator ANC.
Staff has developed such a form, termed an aquatic acidification index (AAI), using a simple
equation to calculate an AAI value in terms of the ambient air indicators NOy and SOx and the
relevant ecological  and atmospheric factors that modify the relationships between the ambient air
indicators and ANC. This AAI reflects the difference between the natural acid neutralizing
capability of a region and acidifying deposition inputs from NOy and SOx in the ambient air.
Recognizing the spatial variability of such factors across the U.S., we conclude it is appropriate
to divide the country into ecologically relevant regions, characterized as acid sensitive or
relatively non-acid  sensitive, and specify the value of each of the factors in the AAI equation for
each such region.
With regard to approaches to defining such ecologically relevant regions, staff concludes that
consideration should be given to using Omernik Ecoregions, level III, as the appropriate  set of
regions over which to define the AAI. There are 84 such level III ecoregions that cover the
continental U.S. This set of ecoregions is based on grouping a variety of vegetation, geological,
and hydrological attributes that are directly relevant to aquatic acidification assessments and that
allow for a practical application of an aquatic acidification standard on a national scale. The
figure below illustrates the Omernik ecoregions with the level III delineations defined by the
different colored areas within each level II group.
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Policy Assessment for the Review of the Secondary National
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                                Omernik Ecoregion II Index Map
          Eco_LevelJII_US
          NA_L2NAME
          I   | ATLANTIC HIGHLANDS

          |   | CBJTRALUSA PLAINS
          |   | CO.D ^EiERTI.

          |   | B/ERGLADES
          |   | MAR INE WEST CO.A — =C =E ST
I   | MI33SSIR=>IALUJVI«-.«NDSOL)n-eASTUSACOASrAL PLAINS I   I TEXA3-LOLISIAkA COASTAL PLAIN

|   | MIXH3 WOOD PLAINS                     |   | UPPER GILA MOUNTAINS

|   | M'.ED '.TOCO SHIELD                     I   I »RM DESERTS

|   | OZARKQUACHITA-APPALAC HIAN FORESTS          |   | WEST-CBJ TRAL SEMI-ARID PRAIRIES

|   | SOUTH CENTRAL SEMI-*fil D PRAIR ES            |   | WESTERN OCR DLLeiA

|   | SO JT4EASTSN LSA =LAIN3                  |   | WESTERN SISRAMADRc P1EJMONT

I   I -SMA.JLPAS-TEXASSEMWRJD FLAW
          |   | MEDITERRAI^SN CALFORNIA |   | TEMPERATE PRAIRIES
With regard to an equation that would define the AAI, staff concludes that consideration should
be given to the following equation, which defines the AAI in terms of four ecological and
atmospheric factors and the ambient air indicators NOy and SOx:

               AAI = Fl - F2 - F3[NOy] - F4[SOX]

In summary, in this equation FI represents the ecosytems natural ability to provide acid
neutralizing capacity and to neutralize nitrogen deposition through plant uptake and other
processes; F2 represents acidifying deposition associated with reduced forms of nitrogen, NHx;
and FS and F4 are the transference ratios that convert concentrations of NOy and SOx to related
deposition of nitrogen and sulfur.  The AAI is constructed from steady state ecosystem modeling,
the atmospheric transference ratios, and incorporation of  reduced froms of nitrogen deposition
(ammonia gas and ammonium ion, expressed as NHx), recognizing that ecosystems respond to
total nitrogen deposition, whether from oxidized or reduced forms of nitrogen.
Factors Fl through F4 would be defined for each ecoregion by specifying ecoregion-specific
values for each factor based on monitored or modeled data that are representative of each
ecoregion.  The Fl factor is also defined by a target ANC value.  More specifically:
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Policy Assessment for the Review of the Secondary National
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    (a) Fl reflects a relative measure of an ecosystem's ability to neutralize acidifying
       deposition.  The value of Fl for each ecoregion would be based on a representative
       critical load for the ecoregion associated with a single national target ANC level, as well
       as a representative runoff rate.  The representative runoff rate, which is also used in
       specifying values for the other factors, would be the median value of the distributions of
       runoff rates within the ecoregion.  The representative critical load would be derived from
       a distribution of critical loads calculated for each water body in the ecoregion for which
       sufficient water quality and hydrology data are available. The representative critical load
       would be defined by selecting a specific percentile of the distribution.
       In identifying a range of percentiles that are appropriate to consider for this purpose, we
       have considered regions characterized as acid sensitive separately from regions
       characterized as relatively non-acid sensitive.  For acid sensitive regions, we conclude
       that consideration should be given to selecting a percentile value from within the range of
       the 70th to the 90th percentile. The lower end of this range was selected to be appreciably
       above the median value so  as to ensure that the critical load would be representative of
       the population of relatively more acid sensitive water bodies within the region, while the
       upper end was selected to avoid the use of a critical load from the extreme tail of the
       distribution which is subject to  a high degree of variability and potential outliers.  For
       relatively non-acid sensitive regions, we conclude that consideration should be given to
       selecting the 50*  percentile to best represent the distribution of water bodies within such
       a region, or alternatively to using the median critical load of all relatively non-acid
       sensitive areas, recognizing that such  areas are far less frequently evaluated than acid
       sensitive areas. Using either of these  approaches would avoid characterizing a generally
       non-acid sensitive region with a critical load that is representative of relatively acid
       sensitive water bodies that  may exist within a generally non-acid sensitive region.
    (b) F2 reflects the deposition of reduced nitrogen. Consideration should be given to
       specifying the value of F2 for each region based on the averaged modeled value across
       the region, using national CMAQ modeling that has been conducted by EPA.
       Consideration could also be given to alternative approaches to specifying this value, such
       as allowance for the use of air quality modeling conducted by States using more refined
       model inputs.
    (c) F3 and F4 reflect transference ratios that convert ambient air concentrations of NOy and
       SOX, respectively, into related deposition of nitrogen and sulfur.  Consideration should be
       given to specifying the values for F3 and F4 for each region based on CMAQ modeling
       results averaged across the region. We conclude that specifying the values for the
       transference ratios based on CMAQ modeling results alone is preferred to an alternative
       approach that combines CMAQ model estimates with observational data.
    (d) The terms [NOy] and [SOX] reflect ambient air concentrations measured at monitoring
       sites within each region.
With regard to averaging time, staff concludes that consideration should be given to averaging
calculated annual AAI values over 3 to 5 years to provide reasonable stability in the resulting
index value, in light of the relatively high degree of interannual variability expected in an index
that is strongly related to the amount and pattern of precipitation that occurs within a region from
year to year.
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With regard to the level of a standard based on the above indicators, alternative forms, and
averaging times, staff concludes that consideration should be given to a level within the range of
20 to 75 |ieq/L.  In reaching this conclusion, staff has considered the available information that
links specific ANC levels to various types of acidification-related effects, and the uncertainties
inherent in such linkages, and the severity of such effects, in sensitive ecosystems, as well as the
extent to which such effects could reasonably be judged to be important from a public welfare
perspective. This range also reflects consideration of the extent to which such a standard would
protect against not only long-term but also episodic acidification, as well as the time lag in
ecosystem response to changes in deposition that may result from such a standard. Relatively
more protection from both long-term and episodic acidification would be provided by a standard
in the mid- to upper part of this range, which would also accelerate the time frame in which the
target ANC level would likely be reached in some sensitive ecosystems. This range also
encompasses target ANC values that have been established by various States and regional and
international organizations to protect against acidification of aquatic ecosystems.
Based on the evidence and assessments in the ISA and REA, we conclude that a target ANC
value of 20 jieq/L is a reasonable lower end of this range, so as to protect against chronic
acidification-related adverse impacts on fish populations which have been characterized as
severe at ANC values below this level.  Further, we conclude that a target ANC value of 75
|ieq/L is a reasonable upper end of this range in recognition that the potential for additional
protection at higher ANC values is substantially more uncertain in light of evidence that
acidification-related effects are far less sensitive to increases in ANC above this value.
As defined above,  an aquatic acidification standard would be interpreted as follows:  the standard
would be met at a monitoring site when the measured annual-average concentrations of NOy and
SOX are such that the value of the annual AAI, averaged over 3 to 5 years, is equal to or greater
than the level of the standard, when using the region-specific values of factors Fl through F4 for
the ecoregion in which the monitor is located.
                                          ES-10

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                          TABLE OF CONTENTS

List of Tables	v
List of Figures	vi
List of Acronyms/Abbreviations	xi
List of Key Terms	xiv

1     INTRODUCTION                                                     1-1
1.1   Purpose	1-1
1.2   Background	1-2
       1.2.1 Legislative Requirements	1-2
       1.2.2 History of Reviews of NAAQS for Nitrogen Oxides and Sulfur Oxides.. 1-3
       1.2.3 History of Related Assessments and Agency Actions	1-6
1.3    Scope of Current Review	1-9
1.4   Conceptual Framework for Deposition-Related Standards	1-11
1.5   Organization of This Document	1-15
1.6   References	1-17

2     CHARACTERIZING EMISSIONS, AIR QUALITY, DEPOSITION AND
      WATER QUALITY	2-1
2.1    Sources of Nitrogen and Sulfur	2-6
      2.1.1 NOX Emissions	2-6
      2.1.2 NH3 Emissions	2-9
      2.1.3 SOX Emissions	2-10
2.2   Ambient Air Characterization	4-14
      2.2.1 Air monitoring networks	2-16
      2.2.2 Overview of CMAQ	2-24
      2.2.3 Overview of air quality using modeled and observed data	2-27
2.3   Characterizing Deposition Through Monitoring and Models	2-34
      2.3.1 Current patterns  of dry and wet deposition	2-36
      2.3.2 Characterizing deposition through CMAQ	2-37
      2.3.3 Relationships between patterns of ambient concentrations and
             deposition	2-40
      2.3.4 Monitoring considerations	2-41
2.4   Characterizing Freshwater Aquatic System Chemistry Using Models and
      Measurements	2-56
      2.4.1 Water chemistry basics	2-56
      2.4.2 Bedrock, soil and vegetation processes relevant aquatic acidification....2-65
      2.4.3 Biogeochemical  ecosystem models used to estimate water quality	2-69
      2.4.4 Water quality network	2-77
2.5   Trends	2-82
      2.5.1 Emissions	2-83
      2.5.2 Air quality and deposition	2-85
      2.5.3 Water quality	2-90
2.6   References	2-96

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3      KNOWN OR ANTICIPATED ECOLOGICAL EFFECTS               3-1
3.1     Acidification: Evidence Of Effects on Structure and Function of Terrestrial
       and Freshwater Ecosystems	3-2
       3.1.1 Nature of acidification-related ecosystem responses	3-3
       3.1.2 Ecosystem sensitivity to acidification	3-12
       3.1.3 Magnitude of ecosystem responses to acidifying deposition	3-12
       3.1.4 Key uncertainties associated with acidification	3-17
3.2     Nitrogen Enrichment: Evidence of Effects on Structure and Function of
       Terrestrial and Freshwater Ecosystems	3-19
       3.2.1 Nature of nutrient enrichment-related ecosystem responses	3-20
       3.2.2 Ecosystem sensitivity to nutrient enrichment	3-22
       3.2.3 Magnitude of ecosystem responses	3-24
       3.2.4 Key uncertainties associated with nutrient enrichment	3-31
3.3     Ecological Effects Associated With Gas-Phase Oxides of Nitrogen
       and Sulfur	3-32
       3.3.1 Nature of ecosystem responses to gas-phase nitrogen and sulfur	3-32
       3.3.2 Ecosystem sensitivity	3-33
       3.3.3 Magnitude of ecosystem responses to gas-phase effects of oxides of
             nitrogen and sulfur	3-34
3.4     Summary	3-34
3.5     References	3-36

4      CONSIDERATIONS OF ADVERSITY TO PUBLIC WELFARE        4-1
4.1     Introduction	4-1
       4.1.1 Benchmarks from Other EPA Programs	4-2
       4.1.2 Other Federal Agencies and the European Union	4-7
4.2     Ecosystem Services and Adversity to Public Welfare	4-10
4.3     Applying Economic Valuation to Ecosystem  Services	4-18
       4.3.1 Economics as aFrameworkto Illustrate Changes in Public Welfare	4-18
       4.3.2 The Role of Economics in Defining Adversity	4-20
       4.3.3 Collective Action as an Indicator of Adversity	4-21
4.4     Effects of Acidification and Nutrient Enrichment on Ecosystem Services	4-25
       4.4.1 Aquatic Acidification	4-26
       4.4.2 Value of Current Ecosystem Service Impairments Due to Aquatic
             Acidification	4-28
       4.4.3 Terrestrial Acidification	4-34
       4.4.4 Value of Current Ecosystem Service Impairments Due to Terrestrial
             Acidification	4-39
       4.4.5 Aquatic Nutrient Enrichment	4-40
       4.4.6 Value of Current Ecosystem Service Impairments Due to Aquatic
             Nutrient Enrichment	4-43
       4.4.7 Terrestrial Nutrient Enrichment	4-44
4.5     References	4-48

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5      CO-PROTECTION FOR OTHER EFFECTS POTENTIALLY
       AFFORDED BY AN AQUATIC ACIDIFICATION STANDARD         5-1
5.1     Potential Co-Protection For Terrestrial Acidification	5-1
5.2     Potential Co-Protection For Terrestrial Nutrient Enrichment	5-4
5.3     Potential Co-Protection For Aquatic Nutrient Enrichment	5-6
5.3     References	5-7

6      ADDRESSING THE ADEQUACY OF THE CURRENT  STANDARDS   6-1
6.1     Appropriateness of the Current Standard	6-1
6.2     Structures of the Current Oxides Of Nitrogen And Sulfur Secondary Standards
       and Relevant Ecological Indicators of Public Welfare Effects	6-3
6.3     Adverse Effects on the Public Welfare Occurring under Current Air Quality
       Conditions for NO2 and SO2	6-9
       6.3.1 Acidification in sensitive aquatic and terrestrial ecosystems	6-11
       6.3.2 Nutrient enrichment effects in sensitive aquatic and terrestrial
             ecosystems	6-16
       6.3.3 Other Ecological Effects Associated with the Deposition of
             Atmospheric Oxides Of Nitrogen and/or Sulfur	6-20
6.4     References	6-22

7      CONSIDERATION OF ALTERNATIVE STANDARDS
       FOR AQUATIC ACIDIFICATION                                     7-1
7.1     Indicators	7-3
       7.1.1 Ambient air indicators for oxides of sulfur	7-4
       7.1.2 Ambient air indicators for oxides of nitrogen	7-4
       7.1.3 Monitoring Considerations	7-7
7.2     Form	7-8
       7.2.1 Ecological indicator	7-10
       7.2.2 Linking ANC to deposition	7-17
       7.2.3 Linking deposition to allowable concentrations	7-21
       7.2.4 Completing the link from ecological indicator to  ambient air
             indicators	7-26
       7.2.5 Spatial Aggregation	7-29
7.3     Averaging Time	7-42
7.4     Level	7-44
7.5     Considerations associated with alternative standards	7-53
7.6     Summary of System Uncertainties	7-64
       7.6.1 Overview	7-64
       7.6.2 Summary of results and  conclusions	7-66
7.7     Summary of Staff Conclusions on Secondary Standards for Oxides of
       Nitrogen and Sulfur	7-75
7.8     References	7-80
                                       in

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                                APPENDICES

Appendix A     Analysis of Critical Loads, Comparing Aquatic and Terrestrial
                Acidification
Appendix B     Methodologies and Assumptions Used In Steady State Ecosystem
                Modeling
Appendix C     Ecoregions, Level III: Description and Summary of Environmental
                Conditions
Appendix D     Maps and Calculation Procedures for Alternative Standards
Appendix E     Derivation To Use Measured Total Nitrate As A Surrogate for NOy
Appendix F      Evaluation of Variability, Sensitivity And Uncertainty In The
                Acidification Index
Appendix G     Cumulative Uncertainty Analysis
                                      IV

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                                 List of Tables
Table 2 -1    Description of parameters, units and conventions	2-4
Table 2-2    Annual National NOx Emissions across Major Source Categories
             in 2002	2-7
Table 2-3a    Annual NOx Emissions across Major Source Categories in 2002 for the
             Eastern United States	2-7
Table 2-3b    Annual NOx Emissions across Major Source Categories in 2002 for the
             Western United States	2-8
Table 2-4    Annual National SO2 Emissions across Major Source Categories in
             2002	2-12
Table 2-5a    Annual SO2 Emissions across Major Source Categories in 2002 for the
             Eastern United States	2-12
Table 2-5b    Annual SO2 Emissions across Major Source Categories in 2002 for the
             Western United States	2-13
Table 2-6    Summary of Monitoring Networks	2-17
Table 2-7    Geographic elements of domains used in RFS2 modeling	2-26
Table 2-8    Summary several commonly used acidification models	2-71
Table 2-9    Summary of data sources  considered for the evaluation of national
             ANC	2-78
Table 2-10    Characteristics of TIME/LTM Sites	2-80
Table 2-11    Period 1 Emissions Density and Period 1-to-Period 3 Relative Changes
             (%) in Oxidized  Sulfur and Nitrogen Emissions, Atmospheric
             Concentration, and Dry, Wet, and Total Deposition	2-85

Table 3-1    Summary of Fish Mortality Response to pH	3-8
Table 3-2    General summary of biological changes anticipated with surface water
             acidification, expressed as a decrease in surface water pH	3-9
Table 3-3    Ecological effects associated with alternative levels of acid neutralizing
             capacity (ANC)	3-15

Table 4-1    Crosswalk between Ecosystem  Services and Public Welfare Effects ... 4-11
Table 4-2    Count of Impacted Lakes	4-29
Table 4-3    Present Value and Annualized Benefits of Recreational Fishing to NY
             Residents, Adirondack Region	4-30
Table 4-4    Aggregate Annual Benefit Estimates of Recreational Fishing to NY
             Residents for the Zero-Out Scenario	4-33
Table 4-5    Annual participation and value  of outdoor (forest related) activity in the
             northeast	4-35
Table 4-6    Summary of Studies of Fall Color Viewing	4-36
Table 4.7    Value Components for WTP for Extensive Protection Program for
             Southern Appalachian Spruce-Fir Forests	4-37
Table 4-8    Summary of WTP Studies for Protection of Spruce Forests in the Southern
             Appalachians	4-38
Table 4-9    Summary of Values for Current Levels of Services and Changes in
             Service Levels in $2007	4-42
                                       v

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Table 4-10    Summary of Annual Damages to Services due to Atmospheric
              Loading	4-43
Table 4-11    Summary of Current Levels of Ecosystem Services	4-46

Table 5.1      Results of the comparison of lake and stream aquatic critical loads (ANC
              of 50 ueq/L) to terrestrial critical loads (Bc:Al molar ratios of 10.0 in soil
              solution)	5-3
Table 5.2      Results of the comparison of lake and stream aquatic critical loads (ANC
              of 50 ueq/L) to terrestrial critical loads (Bc:Al molar ratios of 1.2 in soil
              solution)	5-3

Table 7-5. la.  Calculated AAI values for acid sensitive ecoregions across the range of
              nth percentiles for an alternative level of 20 ueq/L	7-55
Table 7-5. Ib  Calculated AAI values for acid sensitive ecoregions across the range of
              nth percentiles for an alternative level of 35 ueq/L	7-56
Table 7-5. Ic  Calculated AAI values for acid sensitive ecoregions across the range of
              nth percentiles for an alternative level of 50 ueq/L	7-57
Table 7-5. Id  Calculated AAI values for acid sensitive ecoregions across the range of
              nth percentiles for an alternative level of 75 ueq/L	7-58
Table 7-5.2    Summary of the number of acid sensitive ecoregions (out of 29) not likely
              to meet alternative standards based on a 2005 CMAQ simulation	7-59
Table 7-6-1    Summary of Qualitative Uncertainty Analysis of Key Components of the
              AAI	7-70
                                  List of Figures

Figure 1-1  Framework of an ecologically relevant secondary standard to address
           deposition-related effects on sensitive ecosystems	1-12
Figure 1-2  Simplified conceptual design of the form of an aquatic acidification standard
           for oxides of nitrogen and sulfur	1-13

Figure 2-1  Overview of atmospheric, soil and aquatic processes relevant to
           acidification	2-1
Figure 2-2  Diagram illustrating the multiple pollutant, multiple media linkages that are
           incorporated in air quality models like CMAQ	2-2
Figure 2-3  Spatial distribution of annual total NOX emissions (tons/yr) for 2002	2-9
Figure 2-4  Spatial distribution of annual total NHa emissions (tons/yr) for 2002	 2-10
Figure 2-5  Spatial distribution of annual total 862 emissions (tons/yr) for 2002	2-14
Figure 2-6  Routinely operating surface monitoring stations measuring forms of
           atmospheric nitrogen	2-18
Figure 2-7  Routinely operating surface monitoring stations measuring forms of
           atmospheric sulfur	2-19
Figure 2-8  Anticipated network of surface based NOy stations based on 2009 network
           design plans	2-21
                                        VI

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Figure 2-9  Location of NADP passive ammonia sampling locations for the AMON
           network	2-23
Figure 2-10   Map of the CMAQ modeling domain	2-26
Figure 2-11   2005 CMAQ modeled annual average NOY 2-29	
Figure 2-12   2005 CMAQ modeled annual average total reduced nitrogen (NHx)...2-29
Figure 2-13   2005 CMAQ modeled annual average total reduced nitrogen (NH3) ...2-30
Figure 2-14   2005 CMAQ modeled annual average ammonium, NFL;	2-30
Figure 2-15   2005 CMAQ modeled annual average SOX,	2-31
Figure 2-16   2005 CMAQ modeled annual average SO2	2-31
Figure 2-17   2005 CMAQ modeled annual average SO4 	2-32
Figure 2-18   2005 annual average sulfur dioxide concentrations (total mass) based on
             CASTNET	2-32
Figure 2-19   2005 annual average sulfate concentrations (total mass) based on
             CASTNET	2-33
Figure 2-20   Annual average 2005 NOy concentrations from reporting stations in the
             Air Quality System (AQS)	2-33
Figure 2-21   Location of approximately 250 National Atmospheric Deposition
             Monitoring (NADP) National Trends Network (NTN) sites	2-37
Figure 2-22   2005 CMAQ modeled oxidized nitrogen deposition 	2-38
Figure 2-23   2005 CMAQ modeled reduced nitrogen deposition	2-38
Figure 2-24   2005 CMAQ modeled ratio of reduced to total nitrogen deposition	2-39
Figure 2-25   2005 CMAQ modeled oxidized sulfur deposition  	2-39
Figure 2-26   Annual 2002 - 2004 CMAQ derived annual average fraction of ambient
             concentrations and dry deposition of individual NOy species	2-49
Figure 2-27   Examples  of the Relative Abundance of Several NOy Species Measured at
             Two Rural Southeastern Canadian Sites as a Fraction of the Total
             Measured  NOy Concentration	2-50
Figure 2-28   Annual average fraction of NOy ambient air contributed by NO2
             based on 2005 CMAQ Eastern U.S.  simulation at 12 km grid cell
             resolution	2-51
Figure 2-29   Annual average fraction of NOy ambient air contributed by FINOs
             based on 2005 CMAQ Eastern U.S.  simulation at 12 km grid cell
             resolution	2-52
Figure 2-30   Annual average fraction of NOy ambient air contributed by PAN
             based on 2005 CMAQ Eastern U.S.  simulation at 12 km grid cell
             resolution	 2-53
Figure 2-31   Annual 2002 - 2004 CMAQ derived annual average fraction of ambient
             concentrations and dry deposition of individual SOx species	2-54
Figure 2-32   Relationship of the change in total oxidized nitrogen deposition to
             changes in ambient nitric acid and ambient NOy	2-55
Figure 2-33   Equilibrium diagram illustrating distribution of carbonate species as a
             function of pH which is closed to atmospheric CO2 exchange	 2-60
Figure 2-34   The same  system open to atmospheric CO2 exchange where the amount  of
             dissolved carbon is determined by the partial pressure of atmospheric
             CO2andPh	2-60
Figure 2-35   Equilibrium diagram of the log of aluminum species concentrations
                                      vn

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             as a function of pH	2-65
Figure 2-36   Critical loads of acidifying deposition based on MAGIC modeling that
             each surface water location can receive in the Adirondack and
             Shenandoah Case Study Areas	2-73
Figure 2-37   Illustration of a generalized N + S deposition tradeoff curve that is
             calculated by using the FAB approach	2-76
Figure 2-38   Active TIME/LTM sampling locations	 2-79
Figure 2-39   Surface water alkalinity in the conterminous U.S	2-81
Figure 2-40   National map of ANC data (|ieq/L) based on historical and contemporary
             data sorted by ANC classes	2-81
Figure 2-41   Conceptual source to effects pipeline diagram illustrating basic
             accountability concepts as one proceeds from source emissions
             through the air and eventually to effects	2-82
Figure 2-42   Time series trends of all anthropogenic NOx and SOx emissions based
             on EPA's National Emissions Inventory (NEI)	2-84
Figure 2-43   Trends of NOx and SOx reductions associated with EPA's Acid Rain
             Program	2-84
Figure 2-44   Eastern U.S. annual average spatial distribution of SO2 and total nitrate
             concentrations averaged over 1989 -1991  and 2007 - 2009	2-87
Figure 2-45   Eastern U.S. annual average spatial distribution of SO4 concentrations
             averaged over 1989-1991 and 2007-2009	2-88
Figure 2-46   U.S. annual average spatial distribution of wet sulfate and
             nitrate.deposit!on averaged over 1989 -1991 and 2007 - 2009 based on
             theNADP	2-88
Figure 2-47   U.S. annual average spatial distribution of wet ammonium sulfate
             deposition and ambient air ammonium concentrations averaged over
             1989 -1991 and 2007 - 2009 based on CASTNET and NADP	2-89
Figure 2-48   Superimposed Eastern U.S. emission and combined GOME and
             SCIAMACHYNO2 1997-2002 trends	2-89
Figure 2-49   U.S. annual average spatial distribution of hydrogen ion concentration
             in rain water as pH averaged over 1989-1991 and 2007 - 2009	2-90
Figure 2-50   Generalized trends in water quality variables ANC and sulfate ion	2-92
Figure 2-51   Mean rates of change in solute concentration in 16 lakes of the
             Adirondack Long-Term Monitoring (ALTM) program from
             1982 to 2000	2-93
Figure 2-52   Summary of regional trends in surface water chemistry from
             1990 to 2000 	2-93
Figure 2-53   Average NOs" concentrations, SO42" concentrations, and ANC across
             the 44 lakes in the Adirondack Case Study Area	2-94
Figure 2-54   ANC concentrations of preacidification (1860) and 2006 conditions based
             on hindcasts of 44 lakes in the Adirondack Case Study Area modeled
             using MAGIC	2-94
Figure 2-55   Average NOs" concentrations, SO42"concentrations,  and ANC levels for
             the 60 streams in the Shenandoah Case Study Area modeled using
             MAGIC	2-95
                                       Vlll

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Figure 2-56   ANC levels of 1860 (preacidification) and 2006 (current) conditionsbased
             on hindcasts of 60 streams in the Shenandoah Case Study Area modeled
             using MAGIC	2-95

Figure 3-1  Conceptual model of direct and indirect acidification effects on aquatic biota.
           Acidic pollutants (NO3- and SO4-2) lower ANC, resulting in lower pH with
           direct toxic effects on fish	3-4
Figure 3-2  Equilibrium diagram of the log of aluminum species concentrations as a
           function of pH	3-5
Figure 3-3  Mean residual number of species per lake for lakes in Ontario,  by pH
           interval	3-10
Figure 3-4  Benchmarks of atmospheric nitrogen deposition for several ecosystem
           indicators (REA 5.3.1.2) MCF-Mixed Conifer Forest, CSS-Coastal Sage
           Scrub	3-26
Figure 3-5  Observed effects from ambient and experimental atmospheric nitrogen
           deposition loads in relation to using CMAQ 2002 modeling results and
           NADP monitoring data	3-27

Figure 4-1  Common anthropogenic stressors and the essential ecological attributes
           they affect	4-4
Figure 4-2  European maps of eutrophication (left) and acidification (right) which protect
           95% of natural areas in 50x50 km2 European Monitoring and Evaluation
           Programme grid	4-10
Figure 4-3  Representation of the benefits assessment process indicating where some
           ecological benefits may remain unrecognized, unquantified,
           or unmonetized	4-13
Figure 4-4  Conceptual model showing the relationships among ambient air quality
           indicators and exposure pathways and the resulting impacts on ecosystems,
           ecological responses, effects and benefits to characterize known or
           anticipated adverse effects to public welfare	4-15
Figure 4-5. Locations of Eastern U.S. Public Lands relative to deposition of nitrogen
           and sulfur in sensitive aquatic areas	4-16
Figure 4-6. Location of Western U.S. Public Lands relative to deposition of nitrogen and
           sulfur	4-17
Figure 4-7. Conceptual model linking ecological indicator (ANC) to affected ecosystem
           services	4-26

Figure 5.1  Benchmarks of atmospheric nitrogen deposition for several ecosystem
           indicators	5-5

Figure 6-1. Three hour average maximum 2005 SO2 concentrations based on the
           SLAMS reporting to EPA's Air Quality System (AQS) data base	6-7
Figure 6-2  Annual average 2005 NO2 concentrations based on the SLAMS reporting to
           EPA's Air Quality System (AQS) data base	6-8
Figure 6-3  National map highlighting the nine case study areas evaluated
           in the REA	6-12
                                       IX

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Figure 7-1. Conceptual design of the form of an aquatic acidification standard for oxides
           of nitrogen and sulfur	7-9
Figure 7-2.  The relationship between pH and ANC under equilibrium conditions with
           mineral phase gibbsite. Triangles indicate calculated values while circles
           indicate measurements (Bi and Liu 2001)	7-14
Figure 7-3 and 7-4. (Left) Relationship between ANC and number offish species present
           in aquatic freshwater ecosystems in Shenandoah National Park (Source:
           Arthur Bulger, University of Virginia, reproduced from NAPAP, 2005.).
           (Right) Number offish species per lake versus acidity status, expressed as
           ANC, for Adirondack lakes	7-15
Figure 7-5. Omernik Ecoregion II areas  with ecoregion III subdivisions	7-33
Figure 7-6. The top panel captures some of the data details used in delineating acid
           sensitive and relatively non-acid sensitive regions, which are shown in the
           bottom panel	7-36

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AAPI
ADR
A13+
ANC
AQCD
AQRV
                      List of Acronyms and Abbreviations

           Atmospheric Acidification Potential Index
           Adirondack Mountains of New York
           free trivalent aluminum ion
           acid neutralizing capacity
           Air Quality Criteria Document
           air quality related values
ASSETS El Assessment of Estuarine Trophic Status eutrophication index
Bc/Al      Base cation to aluminum ratio, also Be: Al
C          carbon
Ca/Al      calcium to aluminum ratio
Ca2+       calcium ion
CAA      Clean Air Act
CASAC    Clean Air Scientific Advisory Committee
CASTNet  Clean Air Status and Trends Network
CCS       coastal sage scrub
Chi a      chlorophyll a
CLE       critical load exceedance
CMAQ     Community Multiscale Air Quality model
CSS       coastal sage scrub
CWA      Clean Water Act
DIN       dissolved inorganic nitrogen
DO        dissolved oxygen
DOI       U.S. Department of Interior
EMAP     Environmental Monitoring and Assessment Program
EPA       U.S. Environmental Protection Agency
FHWAR   fishing, hunting and wildlife associated recreation survey
FIA       Forest Inventory and Analysis National Program
FWS      Fish and Wildlife Service
GIS       geographic information systems
GPP       gross primary productivity
H+         hydrogen ion
           water vapor
           sulfuric acid
           hectare
           harmful algal bloom
           hydrofluorocarbon
           reactive mercury
           elemental mercury
           nitric acid
HONO     nitrous acid
HUC      hydrologic unit code
IMPROVE Interagency Monitoring of Protected Visual Environments
ISA       Integrated  Science Assessment
K+         potassium ion
H2SO4
ha
HAB
HFC
Hg+2
Hg°
                                       XI

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kg/ha-yr
km
LRMP
LTER
LTM
MAGIC
MCF
MEA
meq/m2-yr
Mg2+
N
N2
N2O
N203
N204
N205
Na+
NAAQS
NADP
NAPAP
NAWQA
NEEA
NEP
NH3
NH4+
(NH4)2SO4
NHx
NO
NO2
NO3"
NOAA
NOX
NOy
NPP
NFS
NRC
NSWS
NTN
NTR
03
OAQPS
OW
PAN
PFC
PH
kilograms per hectare per year
kilometer
Land and Resource Management Plan
Long Term Ecological Monitoring and Research
Long-Term Monitoring
Model of Acidification of Groundwater in Catchments
Mixed Conifer Forest
Millennium Ecosystem Assessment
annual  deposition rate in terms of milliequivalents per square meter
magnesium
nitrogen
gaseous nitrogen
nitrous oxide
nitrogen trioxide
nitrogen tetr oxide
dinitrogen pentoxide
sodium
National Ambient Air Quality Standards
National Atmospheric Deposition Program
National Acid Precipitation Assessment Program
National Water Quality Assessment
National Estuarine Eutrophication Assessment
net ecosystem productivity
ammonia gas
ammonium ion
ammonium sulfate
category label for NH3 plus NH4+
nitric oxide
nitrogen dioxide
 nitrite ion
 nitrate ion
National Oceanic and Atmospheric Administration
nitrogen oxides
total reactive oxidized nitrogen
net primary productivity
National Park Service
National Research Council
National Surface Water Survey
National Trends Network
organic nitrate
ozone
Office of Air Quality Planning and Standards
Office of Water
peroxyacyl nitrates
perfluorocarbons
negative base 10 logarithmic value of hydrogen ion concentration
                                      xn

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ppb
ppm
ppt
PSD
REA
REMAP
S
S203
S207
SAV
SF6
SMP
SO
SO2
SO3
SO32
SO4
SO42
SOM
SOX
2-
2-
parts per billion
parts per million
parts per trillion
prevention of significant deterioration
Risk and Exposure Assessment
Regional Environmental Monitoring and Assessment Program
sulfur
thiosulfate
heptoxide
submerged aquatic vegetation
sulfur hexafluoride
Simple Mass Balance
sulfur monoxide
sulfur dioxide
sulfur trioxide
sulfite ion
wet sulfate
sulfate ion
soil organic matter
sulfur oxides
SPARROW SPAtially Referenced Regressions on Watershed Attributes
SRB       sulfate-reducing bacteria
STORE!   STORage and RETrieval
TIME      Temporally Integrated Monitoring of Ecosystems
TMDL     total maximum daily load
TP         total phosphorus
USFS      U.S. Forest Service
USGS      U.S. Geological Survey
ueq/L      microequivalents per liter
|ig/m3      micrograms per cubic meter
                                      xni

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                                List of Key Terms

Acidification: The process of increasing the acidity of a system (e.g., lake, stream, forest
           soil). Atmospheric deposition of acidic or acidifying compounds can acidify
           lakes, streams, and forest soils.
Air Quality Indicator: The substance or set of substances (e.g., PM2.5, NO2, SO2)
           occurring in the ambient air for which the National Ambient Air Quality
           Standards set a standard level and monitoring occurs.
Alpine: The biogeographic zone made up of slopes above the tree line, characterized by
           the presence of rosette-forming herbaceous plants and low,  shrubby, slow-
           growing woody plants.
Acid Neutralizing Capacity: A key indicator of the ability of water to neutralize the acid
           or acidifying inputs it receives. This ability depends largely on associated
           biogeophysical characteristics, such as underlying geology, base cation
           concentrations, and weathering rates.
Arid Region: A land region of low rainfall, where "low" is widely accepted to be less
           than 250 mm precipitation per year.
Base Cation Saturation: The degree to which soil cation exchange sites are occupied
           with base cations (e.g., Ca2+, Mg2+, K+) as opposed to A13+ and H+. Base
           cation saturation is a measure of soil acidification, with lower values being
           more acidic. There is a threshold whereby soils with base saturations less
           than 20% (especially between 10%-20%) are extremely sensitive to change.
Ecologically Relevant Indicator: A physical, chemical, or biological entity/feature that
           demonstrates a consistent degree of response to a given level of stressor
           exposure and that is easily measured/quantified to make it a useful predictor
           of ecological risk.
Critical Load: A quantitative estimate of an exposure to one or more pollutants, below
           which significant (as defined by the analyst or decision maker) harmful
           effects on specified sensitive elements of the environment do not occur,
           according to present knowledge.
Denitrification: The anaerobic reduction of oxidized nitrogen (e.g., nitrate or nitrite) to
           gaseous nitrogen (e.g., N2O or N2) by denitrifying bacteria.
Dry Deposition: The removal of gases and particles from the atmosphere to surfaces in
           the absence of precipitation (e.g., rain, snow) or occult deposition (e.g., fog).
Ecological Risk: The likelihood that adverse ecological effects may occur or are
           occurring as a result of exposure to one or more stressors (U.S. EPA, 1992).
Ecological Risk Assessment: A process that evaluates the likelihood that adverse
           ecological effects may occur or are occurring as a result of exposure to one
           or more stressors (U.S. EPA, 1992).
Ecosystem: The interactive system formed from all living organisms and their abiotic
           (i.e., physical and chemical) environment within a given area. Ecosystems
           cover a hierarchy of spatial scales and can comprise the entire globe, biomes
           at the continental scale, or small, well-circumscribed systems such as a small
           pond.
Ecosystem Benefit: The value,  expressed qualitatively, quantitatively, and/or in
           economic terms, where possible, associated with changes in ecosystem
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           services that result either directly or indirectly in improved human health
           and/or welfare. Examples of ecosystem benefits
           that derive from improved air quality include improvements in habitats for
           sport fish species, the quality of drinking water and recreational areas, and
           visibility.
Ecosystem Function: The processes and interactions that operate within an ecosystem.
Ecosystem Services: The ecological processes or functions having monetary or non-
           monetary value to individuals or society at large. These are (1) supporting
           services, such as productivity or biodiversity maintenance; (2) provisioning
           services, such as food, fiber, or fish; (3) regulating services,  such as climate
           regulation or carbon sequestration; and (4) cultural services, such  as tourism
           or spiritual and aesthetic appreciation.
Equivalents. Concentrations in terms of the electrical charge units associated with ionic
           charge in water or the potential of atmospheric deposition of N or S to
           produce ions.  Coonmly reported as |ieq/l (water concentration) and meq/m2-
           yr (deposition)
Eutrophication: The process by which nitrogen additions stimulate the growth of
           autotrophic biota, usually resulting in the depletion of dissolved oxygen.
Nitrification: The oxidation through plant, soil and microbiologicak processes of
           reduced nitrogen (ammonia gas and ammonium ion) into nitrite and evetually
           nitrate.
Nitrogen Enrichment: The process by which a terrestrial system becomes enhanced by
           nutrient additions to a degree that stimulates the growth of plant or other
           terrestrial biota, usually resulting in an increase in productivity.
Nitrogen Saturation: The point at which nitrogen inputs from atmospheric deposition
           and other sources exceed the biological requirements of the ecosystem; a
           level beyond nitrogen enrichment.
Occult Deposition: The removal of gases and particles from the atmosphere to surfaces
           by fog or mist.
Semi-arid Regions: Regions of moderately low rainfall, which are not highly productive
           and are usually classified as rangelands. "Moderately low" is widely
           accepted as between 100- and 250-mm precipitation per year.
Sensitivity: The degree to which a system is affected, either adversely or beneficially, by
           NOx and/or SOx pollution (e.g., acidification, nutrient enrichment). The
           effect may be direct (e.g., a change in growth in response to a change in  the
           mean, range, or variability of nitrogen deposition) or indirect (e.g., changes
           in growth due to the direct effect of nitrogen consequently altering
           competitive dynamics between species and decreased biodiversity).
Total Reactive Nitrogen: This includes all biologically, chemically, and radiatively
           active nitrogen compounds in the atmosphere and biosphere, such as NH3,
           NH4+, NO, NO2, HNO3, N2O, NO3-, and organic compounds (e.g., urea,
           amines, nucleic acids).
Valuation: The economic or non-economic process of determining either the value of
           maintaining a given ecosystem type, state, or condition, or the value of a
           change in an ecosystem, its components, or the services it provides.
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Variable Factors: Influences which by themselves or in combination with other factors
           may alter the effects on public welfare of an air pollutant (section 108 (a)(2))
       (a) Atmospheric Factors: Atmospheric conditions that may influence
           transformation, conversion, transport, and deposition, and thereby, the effects
           of an air pollutant on public welfare, such as precipitation, relative humidity,
           oxidation state, and co-pollutants present in the atmosphere.
       (b) Ecological Factors: Ecological conditions that may influence the effects of an
           air pollutant on public welfare once it is introduced into an ecosystem, such
           as soil base saturation, soil thickness, runoff rate, land use conditions,
           bedrock geology, and weathering rates.
Vulnerability: The degree to which a system is susceptible to, and unable to cope with,
           the adverse effects of NOx and/or SOx air pollution.
Welfare Effects: The effects on soils, water, crops, vegetation, man-made materials,
           animals, wildlife, weather, visibility, and climate; as well as damage to and
           deterioration of property, hazards to transportation, and the effects on
           economic values and on personal comfort and well-being, whether caused by
           transformation, conversion, or combination with other air pollutants (Clean
           Air Act Section 302[h]).
Wet Deposition: The removal of gases and particles from the atmosphere to surfaces by
           rain or other precipitation.
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                                1   INTRODUCTION
1.1    PURPOSE
       The U.S. Environmental Protection Agency (EPA) is presently conducting a joint review
of the secondary (welfare-based) national ambient air quality standards (NAAQS) for oxides of
nitrogen and oxides of sulfur.  The EPA's overall plan and schedule for this review were
presented in the Integrated Review Plan for the Secondary National Ambient Air Quality
Standards for Nitrogen Dioxide and Sulfur Dioxide (U.S. EPA, 2007). The IRP identified key
policy-relevant issues to be addressed in this review as a series of questions that frame our
consideration of whether the current secondary NAAQS for oxides of nitrogen and sulfur should
be retained or revised.
       This Policy Assessment (PA), prepared by staff in the EPA's Office of Air Quality
Planning and Standards (OAQPS) is intended to help "bridge the gap" between the relevant
scientific information and assessments and the judgments required of the EPA Administrator in
determining whether, and if so how, it is appropriate to revise the secondary NAAQS for oxides
of nitrogen and sulfur. *  This PA presents factors relevant to EPA's review of the secondary
NAAQS for oxides of nitrogen and oxides of sulfur. It focuses on both evidence- and risk-based
information, together with related limitations and uncertainties, in evaluating the adequacy of the
current NAAQS and in identifying potential  alternative standards for  consideration.
       In this PA, we consider the policy implications of the scientific information available in
this review as assessed in the Integrated Science Assessment for Oxides of Nitrogen and Sulfur-
Ecological Criteria (U.S. EPA, 2008), prepared by EPA's National Center for Environmental
Assessment (NCEA), and the results from the analyses contained in the Risk and Exposure
Assessment for Review of the Secondary National Ambient Air Quality Standards for Oxides of
Nitrogen and Oxides of Sulfur (U.S. EPA, 2009).2 In so doing, we focus on information that is
most pertinent to evaluating the basic elements of NAAQS:  indicator3, averaging time, form,4
1 Preparation of a PA by OAQPS staff reflects Administrator Jackson's decision to modify the NAAQS review
process that was presented in the IRP. See http://www.epa.gov/ttn/naaqs/review.html for more information on the
current NAAQS review process.
2 These documents are available at http://www.epa.gov/ttn/naaqs/standards/no2so2sec/index.html.
3 The "indicator" of a standard defines the chemical species or mixture that is to be measured in determining
whether an area attains the standard.
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and level. These elements, which together serve to define each standard, must be considered
collectively in evaluating the public welfare protection afforded by these standards.
       Although this PA should be of use to all parties interested in this review of the secondary
NAAQS for oxides of nitrogen and sulfur, it is written with an expectation that the reader has
some familiarity with the technical discussions contained in the ISA and REA.

1.2    BACKGROUND
1.2.1  Legislative Requirements
       Two sections of the Clean Air Act (CAA) govern the establishment and revision of the
NAAQS. Section 108 (42 U.S.C. section 7408) directs the Administrator to identify and list
certain air pollutants and then to issue air quality criteria for those pollutants. The Administrator
is to list those "air pollutant[s]...  [that] in his judgment, cause or contribute to air pollution
which may reasonably be anticipated to endanger public health or welfare; the presence of which
in the ambient air results from numerous or diverse mobile or stationary sources; [and] for which
air quality criteria had not been listed prior to December 31, 1970, but for which [the
Administrator] plans to issue air quality criteria...".  Air quality criteria are intended to
"accurately  reflect the latest scientific knowledge useful in indicating the kind and extent of all
identifiable  effects on public health or welfare which may be expected from the presence of [a]
pollutant in  the ambient air ...."  The air quality criteria include "(A) those variable factors
(including atmospheric conditions) which of themselves or in combination with other factors
may alter the effects on public health or welfare of such air pollutant; (B) the types of air
pollutants which, when present in the atmosphere, may interact with such pollutant to produce an
adverse effect on public health of welfare; and (C) any known or anticipated adverse effects on
welfare." 42 U.S.C. § 7408(b).
       Section 109 (42 U.S.C. section 7409) directs  the Administrator to propose and
promulgate  "primary" and "secondary" NAAQS for  pollutants  for which air quality criteria are
issued.  Section 109(b)(l) defines a primary standard as one "the attainment and maintenance of
which in the judgment of the Administrator, based on [air quality] criteria and allowing an
adequate margin of safety, are requisite to protect the public health." A secondary standard, as
4 The "form" of a standard defines the metric that is to be compared to the level of the standard in determining
whether an area attains the standard.
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defined in section 109(b)(2), must "specify a level of air quality the attainment and maintenance
of which, in the judgment of the Administrator, based on such criteria, is requisite to protect the
public welfare from any known or anticipated adverse effects associated with the presence of
such air pollutant in the ambient air."5
       In setting standards that are "requisite" to protect public health and welfare, as provided
in section 109(b), EPA's task is to establish standards that are neither more nor less stringent
than necessary for these purposes. Whitman v. American Trucking Associations, 531 U.S. 457,
473. In establishing "requisite" EPA may not consider the costs of implementing the standards.
Id. at 471.  Likewise, "[attainability and technological feasibility are not relevant considerations
in the promulgation of national ambient air quality standards." American Petroleum Institute v.
CostlejS65F. 2dat 1185.
       Section 109(d) (1) of the CAA requires that "[n]ot later than December 31, 1980,  and at
5-year intervals thereafter, the Administrator shall complete a thorough review of the criteria
published under section 108 and the national ambient air quality standards . .  . and shall make
such revisions in such criteria and standards and promulgate such new standards as may be
appropriate ..." 42 U.S.C. § 7409(d)(l). Section 109(d)(2) requires that an independent scientific
review committee "shall  complete a review of the criteria  ... and the national primary and
secondary ambient  air quality standards ... and shall recommend to the Administrator any new ...
standards and revisions of existing criteria and standards as may be appropriate ...."  42 U.S.C. §
7409(d)(2). This independent review function is performed by the Clean Air  Scientific Advisory
Committee (CAS AC) of EPA's Science Advisory Board.
1.2.2  History of Reviews of NAAQS for Nitrogen Oxides and Sulfur Oxides
NAAQS for Oxides of Nitrogen
       After reviewing the relevant science on the public health and welfare effects associated
with oxides of nitrogen, EPA promulgated identical primary and secondary NAAQS for NC>2 in
April 1971. These  standards wer set at a level of 0.053 parts per million (ppm) as an annual
average (36 FR 8186). In 1982,  EPA published Air Quality Criteria for Oxides of Nitrogen (U.S.
5 Welfare effects as defined in section 302(h) (42 U.S.C. section 7602(h)) include, but are not limited to, "effects on
soils, water, crops, vegetation, man-made materials, animals, wildlife, weather, visibility and climate, damage to and
deterioration of property, and hazards to transportation, as well as effects on economic values and on personal
comfort and well-being."
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EPA, 1982), which updated the scientific criteria upon which the initial standards were based. In
February 1984 EPA proposed to retain these standards (49 FR 6866). After taking into account
public comments, EPA published the final decision to retain these standards in June 1985 (50 FR
25532).
       The EPA began the most recent previous review of the oxides of nitrogen secondary
standards in 1987. In November 1991  EPA released an updated draft AQCD for CAS AC and
public review and comment (56 FR 59285), which provided a comprehensive assessment of the
available scientific and technical information on health and welfare effects associated with NCh
and other oxides of nitrogen.  The CAS AC reviewed the draft document at a meeting held on
July 1, 1993, and concluded in a closure letter to the Administrator that the document "provides a
scientifically balanced and defensible summary of current knowledge of the effects of this
pollutant and provides an adequate basis for EPA to make a decision as to the appropriate
NAAQS for NCh" (Wolff, 1993).  The Air Quality Criteria for Oxides of Nitrogen was then
finalized (U.S. EPA, 1993). EPA's OAQPS also prepared a Staff Paper that summarized and
integrated the key studies and scientific evidence contained in the revised AQCD for oxides of
nitrogen and identified the critical elements to be considered in the review of the NCh NAAQS.
CAS AC reviewed two drafts of the Staff Paper and concluded in a closure letter to the
Administrator that the document provided a "scientifically adequate basis for regulatory
decisions on nitrogen dioxide" (Wolff,  1995).
       In October 1995 the Administrator announced her  proposed decision not to revise either
the primary or secondary NAAQS for NCh (60 FR 52874; October 11, 1995). A year later, the
Administrator made a final determination not to revise the NAAQS for NO2 after careful
evaluation of the comments received on the proposal (61 FR 52852; October 8, 1996). While the
primary NO2 standard was revised in January 2010 by  supplementing the existing annual
standard with the establishment of a new 1-hour standard (75 FR 6474), the secondary NAAQS
for NO2 remains 0.053 ppm (100 micrograms per cubic meter [ug/m3] of air), annual arithmetic
average, calculated as the arithmetic mean of the 1-hour NO2 concentrations.
NAAOS for Oxides of Sulfur
       EPA promulgated primary and  secondary NAAQS for SO2 in April 1971 (36 FR 8186).
The secondary standards included a standard set at 0.02 ppm, annual arithmetic mean, and a 3-
hour average standard set at 0.5 ppm, not to be exceeded more than once per year.  These
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secondary standards were established solely on the basis of evidence of adverse effects on
vegetation. In 1973, revisions made to Chapter 5 ("Effects of Sulfur Oxide in the Atmosphere on
Vegetation") of Air Quality Criteria for Sulfur Oxides (U.S. EPA, 1973) indicated that it could
not properly be concluded that the vegetation injury reported resulted from the average 862
exposure over the growing season, rather than from short-term peak concentrations.  Therefore,
EPA proposed (38 FR 11355) and then finalized (38 FR 25678) a revocation of the annual mean
secondary standard. At that time, EPA was aware that then-current concentrations of oxides of
sulfur in the ambient air had other public welfare effects, including effects on materials,
visibility, soils, and water. However, the available data were considered insufficient to establish
a quantitative relationship between specific ambient  concentrations of oxides of sulfur and such
public welfare effects (38 FR 25679).
       In 1979, EPA announced that it was revising the AQCD for oxides of sulfur concurrently
with that for paniculate matter (PM) and would produce a combined PM and oxides of sulfur
criteria document.  Following its review of a draft revised criteria document in August 1980,
CAS AC concluded that acid deposition was a topic of extreme scientific complexity because of
the difficulty in establishing firm quantitative relationships among (1) emissions of relevant
pollutants (e.g., SC>2 and oxides of nitrogen), (2) formation of acidic wet and dry deposition
products, and (3) effects on terrestrial and aquatic ecosystems. CASAC also noted that acid
deposition involves, at a minimum,  several different criteria pollutants:  oxides of sulfur, oxides
of nitrogen, and the fine particulate fraction of suspended particles.  CASAC felt that any
document on this subject should address both wet and dry deposition, since dry deposition was
believed to account for a substantial portion of the total  acid deposition problem.
       For these reasons, CASAC recommended that a  separate, comprehensive document on
acid deposition be prepared prior to any consideration of using the NAAQS as a regulatory
mechanism for the control of acid deposition. CASAC also suggested that a discussion of acid
deposition be included in the AQCDs for oxides of nitrogen and PM and oxides of sulfur.
Following CASAC closure on the AQCD for oxides of sulfur in December 1981, EPA's OAQPS
published a Staff Paper in November 1982, although the paper did not directly assess the issue of
acid deposition. Instead, EPA subsequently prepared the following documents to address acid
depo sition: The Acidic Deposition Phenomenon and Its Effects: Critical Assessment Review
Papers, Volumes Iand II(U.S. EPA, 1984a, b) and The Acidic Deposition Phenomenon and Its

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Effects: Critical Assessment Document (U.S. EPA, 1985) (53 FR 14935 -14936).  These
documents, though they were not considered criteria documents and did not undergo CAS AC
review, represented the most comprehensive summary of scientific information relevant to acid
deposition completed by EPA at that point.
       In April 1988 (53 FR 14926), EPA proposed not to revise the existing primary and
secondary standards for 862. This proposed decision with regard to the secondary 862 NAAQS
was due to the Administrator's conclusions that (1) based upon the then-current scientific
understanding of the acid deposition problem, it would be premature and unwise to prescribe any
regulatory control program at that time and (2) when the fundamental scientific uncertainties had
been decreased through ongoing research efforts, EPA would draft and support an appropriate set
of control measures.  Although EPA revised the primary SC>2 standard in June 2010 by
establishing a new 1-hour standard and revoking the existing 24-hour and annual standards (75
FR 35520), no further decisions on the secondary 862 standard have been published.
1.2.3   History of Related Assessments and Agency Actions
       In 1980, the Congress created the National Acid Precipitation Assessment Program
(NAPAP) in response to growing concern about acidic deposition.  The NAPAP was given a
broad 10-year mandate to examine the causes and effects of acidic deposition and to explore
alternative control options to alleviate acidic deposition and its effects. During the course of the
program, the NAPAP issued a series of publicly available interim reports prior to the completion
of a final report in 1990 (NAPAP, 1990).
       In spite of the complexities and significant remaining uncertainties associated with the
acid deposition problem, it soon became clear that a program to address acid deposition was
needed. The Clean Air Act Amendments of 1990 included numerous separate provisions related
to the acid deposition problem.  The primary and most important of the provisions, the
amendments to Title IV of the Act, established the Acid Rain Program to reduce emissions of
SO2 by 10 million tons and emissions of nitrogen oxides by 2 million tons from 1980 emission
levels in order to achieve reductions over broad geographic regions. In this provision, Congress
included a statement of findings that led them to take action, concluding that (1) the presence of
acid compounds and their precursors in the atmosphere and in deposition from the atmosphere
represents a threat to natural resources, ecosystems, materials, visibility, and public health; (2)
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the problem of acid deposition is of national and international significance; and (3) current and
future generations of Americans will be adversely affected by delaying measures to remedy the
problem.
       Second, Congress authorized the continuation of the NAPAP in order to assure that the
research and monitoring efforts already undertaken would continue to be coordinated and would
provide the basis for an impartial assessment of the effectiveness of the Title IV program.
       Third, Congress considered that further action might be necessary in the long term to
address any problems remaining after implementation of the Title IV program and, reserving
judgment on the form that action could take, included Section 404 of the 1990 Amendments
(Clean Air Act Amendments of 1990, Pub. L. 101-549,  § 404) requiring EPA to conduct a study
on the feasibility and effectiveness of an acid deposition standard or standards to  protect
"sensitive and critically sensitive aquatic and terrestrial resources."  At the conclusion of the
study, EPA was to submit a report to Congress. Five years later, EPA submitted  its report,
entitled Acid Deposition Standard Feasibility Study:  Report to Congress (U.S. EPA, 1995) in
fulfillment of this requirement. That Report concluded that establishing acid deposition
standards for sulfur and nitrogen deposition may at some point in the future be technically
feasible, although appropriate deposition loads for these acidifying chemicals could not be
defined with reasonable certainty at that time.
       Fourth, the 1990 Amendments also added new language to sections of the CAA
pertaining to the scope and application of the secondary NAAQS  designed to protect the public
welfare.  Specifically, the definition of "effects on welfare" in Section 302(h) was expanded to
state that the welfare effects include effects ".. .whether caused by transformation, conversion, or
combination with other air pollutants."
       In 1999, seven Northeastern states cited this amended language in Section 302(h) in a
petition asking EPA to use its authority under the NAAQS program to promulgate secondary
NAAQS for the criteria pollutants associated with the formation of acid rain. The petition stated
that this language "clearly references the transformation of pollutants resulting in the inevitable
formation of sulfate and nitrate aerosols and/or their  ultimate environmental impacts as wet and
dry deposition, clearly signaling Congressional intent that the welfare damage occasioned by
sulfur and nitrogen oxides be addressed through the secondary standard provisions of Section
109 of the Act."  The petition further stated that "recent federal studies, including the NAPAP

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Biennial Report to Congress: An Integrated Assessment, document the continued and increasing
damage being inflicted by acid deposition to the lakes and forests of New York, New England
and other parts of our nation, demonstrating that the Title IV program had proven insufficient."
The petition also listed other adverse welfare effects associated with the transformation of these
criteria pollutants, including impaired visibility, eutrophication of coastal estuaries, global
warming, and tropospheric ozone and stratospheric ozone depletion.
       In a related matter, the Office of the Secretary of the U.S. Department of Interior (DOI)
requested in 2000 that EPA initiate a rulemaking proceeding to enhance the air quality in
national parks and wilderness areas in order to protect resources and values that are being
adversely affected by air pollution. Included among the effects of concern identified in the
request were the acidification of streams, surface waters, and/or soils; eutrophication of coastal
waters; visibility impairment; and foliar injury from ozone.
       In a Federal Register notice in 2001, EPA announced receipt of these requests and asked
for comment on the issues raised in them. EPA stated that it would consider any relevant
comments and information submitted, along with the information provided by the petitioners and
DOI, before making any decision concerning a response to these requests for rulemaking (65 FR
48699).
       The 2005 NAPAP report states that "... scientific studies indicate that the emission
reductions achieved by Title IV are not sufficient to allow recovery of acid-sensitive ecosystems.
Estimates from the literature of the scope of additional emission reductions that are necessary in
order to protect acid-sensitive ecosystems range from approximately 40-80% beyond full
implementation of Title IV...." The results of the modeling presented in this Report to Congress
indicate that broader  recovery is not predicted without additional emission reductions" (NSTC,
2010).
       Given the state of the science as described in the ISA and in other recent reports, such as
the NAPAP reports noted above, EPA has decided, in the context of evaluating the adequacy of
the current NO2 and SO2 secondary standards in this review, to revisit the question of the
appropriateness of setting secondary NAAQS to address remaining known or anticipated adverse
public welfare effects resulting from the acidic  and nutrient deposition of these criteria
pollutants.
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1.3    SCOPE OF CURRENT REVIEW
       In conducting this periodic review of the secondary NAAQS for oxides of nitrogen and
oxides of sulfur, as discussed in the IRP, EPA decided to assess the scientific information,
associated risks, and standards relevant to protecting the public welfare from adverse effects
associated jointly with oxides of nitrogen and sulfur.  Although EPA has historically adopted
separate secondary standards for oxides of nitrogen and oxides of sulfur, EPA is conducting a
joint review of these standards because oxides of nitrogen and sulfur, and their associated
transformation products are linked from an atmospheric chemistry perspective, as well as from
an environmental effects perspective.  The National Research Council (NRC) has recommended
that EPA  consider multiple pollutants, as appropriate, in forming the scientific basis for the
NAAQS (NRC, 2004). As discussed in the ISA and REA, there is a strong basis for considering
these pollutants together, building upon EPA's past recognition of the interactions of these
pollutants and on the growing body of scientific information that is now available related to these
interactions and associated ecological effects.
       In defining the scope of this review, we recognize that EPA has set secondary standards
for two other criteria pollutants related to oxides of nitrogen and sulfur:  ozone and particulate
matter (PM). Oxides of nitrogen are precursors to the formation of ozone in the atmosphere, and
under certain conditions, can combine with atmospheric ammonia to form ammonium nitrate, a
component of fine PM.  Oxides of sulfur are precursors to the  formation of particulate sulfate,
which is a significant component of fine PM in many parts of the U.S.  There are a number of
welfare effects directly associated with ozone and  fine PM, including ozone-related damage to
vegetation and PM-related visibility impairment. Protection against those effects is provided by
the ozone and fine PM secondary  standards.  This  review focuses on evaluation of the protection
provided by secondary standards for oxides of nitrogen and sulfur for two general types of
effects: (1) direct effects on vegetation associated with exposure to gaseous oxides of nitrogen
and sulfur in the ambient air, which are the effects that the current NO2 and SO2 secondary
standards  protect against and (2) effects associated with the deposition of oxides of nitrogen and
sulfur to sensitive aquatic and terrestrial ecosystems, including deposition in the form of
particulate nitrate and particulate sulfate.
       The ISA focuses on the ecological effects associated with deposition of ambient oxides of
nitrogen and sulfur to natural sensitive ecosystems, as distinguished from commercially managed
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forests and agricultural lands. This focus reflects the fact that the majority of the scientific
evidence regarding acidification and nutrient enrichment is based on studies in unmanaged
ecosystems. Non-managed terrestrial ecosystems tend to have a higher fraction of N deposition
resulting from atmospheric nitrogen (ISA, section 3.3.2.5). In addition, the ISA notes that
agricultural and commercial forest lands are routinely fertilized with amounts of nitrogen that
exceed air pollutant inputs even in the most polluted areas (ISA, section 3.3.9).  This review
recognizes that the effects of nitrogen deposition in managed areas are viewed differently from a
public welfare perspective than  are the effects of nitrogen deposition in natural, unmanaged
ecosystems, largely due to the more homogeneous,  controlled nature of species composition and
development in managed ecosystems and the potential for benefits of increased productivity in
those ecosystems.
       In focusing on natural sensitive ecosystems, this PA primarily considers the effects of
ambient oxides of nitrogen and sulfur via deposition on multiple ecological receptors. The ISA
highlighted effects including those associated with acidification and nitrogen nutrient
enrichment. With a focus on these deposition-related effects, EPA's objective is to develop a
framework for oxides of nitrogen  and sulfur standards that incorporates ecologically relevant
factors and that recognizes the interactions between the two pollutants as they deposit to
sensitive ecosystems.  The overarching policy objective is to develop a secondary standard(s)
that is based on the ecological criteria described in the ISA and the results of the assessments in
the REA, and is consistent with the requirement of the CAA to set secondary standards that are
requisite to protect the public welfare from any known or anticipated adverse effects  associated
with the presence of these air pollutants in the ambient air.  Also consistent with the CAA, as
discussed above in section 1.2.1, this policy objective necessarily includes consideration of
"variable factors . . . which of themselves or in combination with other factors may alter the
effects on public welfare" of the criteria air pollutants included in this review.
       In addition, we have chosen to focus on the  effects of ambient oxides of nitrogen and
sulfur on ecological impacts on senstitive aquatic ecosystems associated with acidifying
deposition of nitrogen and sulfur,  which is a transformation product of ambient oxides of
nitrogen and sulfur. Based on the information in the ISA, the assessments presented  in the REA,
and advice from CAS AC on earlier drafts of this PA (Russell and Samet, 2010a, 201 Ob), and as
discussed below in chapter 2, we have the greatest confidence in the causal  linkages between

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oxides of nitrogen and sulfur and aquatic acidification effects relative to other deposition-related
effects, including terrestrial acidification and aquatic and terrestrial nutrient enrichment.
       In developing policy options for the Administrator's consideration, we note that decisions
on retaining or revising the current secondary standards for oxides of nitrogen and sulfur are
largely public welfare policy judgments based on the Administrator's informed assessment of
what constitutes requisite protection against adverse effects to public welfare. A public welfare
policy decision should draw upon scientific information and analyses about welfare effects,
exposure and risks, as well as judgments about the appropriate response to the range of
uncertainties that are inherent in the scientific evidence and analyses. The ultimate determination
as to what level of damage to ecosystems and the services provided by those ecosystems is
adverse to public welfare is not wholly a scientific question, although it is informed by scientific
studies linking ecosystem damage to losses in ecosystem services, and information on the value
of those losses of ecosystem services. In reaching such decisions, the Administrator seeks to
establish standards that are neither more nor less stringent than necessary for this purpose.

1.4    CONCEPTUAL FRAMEWORK FOR DEPOSITON-RELATED STANDARDS
       As noted above, there is a strong basis for considering deposition-related standards for
oxides of nitrogen and sulfur together at this time, building upon EPA's and CASAC's
recognition of the interactions of these pollutants and on the growing body of scientific
information that is now available related to these interactions and associated ecological effects.
The REA introduced a conceptual framework for ecologically meaningful secondary standards
that recognized the complex processes by which ecosystems are exposed through deposition to
ambient oxides of nitrogen and sulfur.  That framework provided a flow from ambient
concentrations to exposures via deposition to ecological indicators and effects (Figure ES-2 in
the REA Executive Summary). Figure 1-1 below is an adaptation of the REA framework, which
represents the process by which we can determine the deposition-related risks to sensitive
aquatic and terrestrial ecosystems associated with ambient concentrations of oxides of nitrogen
and sulfur.  This framework illustrates how a level of protection related to an indicator of
ecological effect(s) can be linked to atmospheric concentrations of indicators of oxides of
nitrogen and sulfur. It illustrates the linkages between ambient air concentrations and resulting
deposition metrics, and between the deposition metric and the ecological indicator of concern.
                                       1-11

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What is referred to as an atmospheric deposition transformation function translates ambient
atmospheric concentrations of oxides of nitrogen and sulfur to nitrogen and sulfur deposition
metrics, while an ecological effect function transforms the deposition metric into an ecological
indicator.
      Structure  of an  Ecologically-based Standard
                   Variable/Fixed
                     Factors:
                    Atmospheric
                    Landscape
                   Atmospheric
                    Deposition
                   Transformation
                     Function
Deposition
 Metric
                                     T
                               Form of the Standard
                               Level of the Standard
Figure 1-1.  Framework of an ecologically relevant secondary standard to address deposition-
related effects on sensitive ecosystems (adapted from the REA, Figure ES-2).

       Development of a form for an ecologically relevant standard that reflects this structure is
a critical step in the overall standard setting process. The atmospheric levels of oxides of
nitrogen and sulfur that afford a particular level of ecosystem protection are those levels that
result in an amount of deposition that is less than the amount of deposition that a given
ecosystem can accept without defined levels of degradation of the ecological indicator for a
targeted effect.
       Drawing from the framework developed in the REA, the framework we are using to
structure an ecologically meaningful secondary standard in this PA is depicted below in a more
                                       1-12

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simple graphic, Figure 1-2, that highlights the three key linkages that need to be considered in
developing an ecologically relevant standard.
                         Linking atmospheric
                         oxides of S and N deposition to
                         ecological indicator
  Ecological effects and                                    Linking deposition to "allowable"
  ecological indicator                                      concentrations of ambient air
                                                        indicators of oxides of N and S

Figure 1-2. Simplified conceptual design of the form of an aquatic acidification standard for
oxides of nitrogen and sulfur.
       The details of this simplified conceptual framework are discussed in chapter 7, including
discussions of modifying factors that alter the relationship between ambient atmospheric
concentrations of oxides of nitrogen and sulfur and depositional loads of nitrogen and sulfur, and
those that modify the relationship between deposition loads and the ecological indicator.
       In setting NAAQS to protect public health and welfare, EPA has historically established
standards which require the comparison of monitored concentrations of an air pollutant against a
numerical metric of atmospheric concentration that does not vary geographically. This approach
has appropriately protected public health, as at-risk populations are widely distributed throughout
the nation. As more is learned about the effects of pollutants such as oxides of nitrogen and
sulfur and the environment, however, such an approach cannot effectively or consistently
provide the requisite level of protection to public welfare from effects on sensitive ecosystems.
In this review, we are considering a standard that takes into account variable factors, such as
atmospheric variables and location-specific characteristics of ecosystems, as the appropriate
approach to protect the public welfare from the effects associated with the presence of these
pollutants in the ambient air.
       While EPA has most often considered the results of direct exposure to an air pollutant in
the ambient air in assessing effects on public health and welfare, such as the health effects on
humans when breathing in  an air pollutant or the effects on vegetation through the uptake of air
pollutants from the ambient air through leaves, EPA has also considered, where appropriate, the
                                        1-13

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effects of exposure to air pollutants through more indirect mechanisms. For example, both in
1978 and in 2008, EPA established a NAAQS for lead that addressed the health effects of
ambient lead whether the lead particles were inhaled or were ingested after deposition on the
ground or other surfaces. 73 FR 66964 (November 12, 2008), Lead Industries v. EPA, 647 F.2d
1130 (DC Cir. 1980) (1978 NAAQS).  The deposition of ambient oxides of nitrogen and sulfur
to terrestrial and aquatic environments can impact ecosystems through both direct and indirect
mechanisms, as discussed in the REA and below in chapter 2.  Given Congress' instruction to set
a standard that "is requisite to protect the public welfare from "any known or anticipated adverse
effects associated with the presence of such air pollutant in the ambient air," 42 U.S.C. § 109
(b)(2), this review appropriately considers widely acknowledged effects, such as acidification
and nutrient enrichment, which are associated with the presence of oxides of nitrogen and sulfur
in the ambient air through the deposition of nitrogen and  sulfur that results from oxides of
nitrogen and sulfur in the ambient air.
       In this review, we are considering the development of a standard that takes into account
the variability in deposition-related effects associated with levels of oxides of nitrogen and sulfur
in the ambient air.  The CAA requires EPA to establish "national" standards, based on the air
quality criteria that provide the requisite degree of protection, but does not clearly address how to
do so under the circumstances present here.  In this PA we develop an approach that is designed
to provide a generally uniform degree of protection throughout the country by allowing for
varying concentrations of allowable ambient oxides of nitrogen and sulfur, depending on
atmospheric conditions and other ecological variables, to achieve that degree of protection.
Such a standard would protect sensitive ecosystems wherever such ecosystems  are found.  This
approach recognizes that setting a standard that is sufficient to  protect the public welfare, but not
more than is necessary, calls for consideration of a standard such as the one discussed in this
document.

 1.5    ORGANIZATION OF THIS DOCUMENT
       This PA includes staffs evaluation of the policy implications  of the scientific assessment
of the evidence presented and assessed in the ISA and the results of quantitative assessments
based on that information presented and assessed in the REA.  Taken together,  this information
informs staff conclusions and the identification of policy  options for consideration in addressing
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public and welfare effects associated with the presence of oxides of nitrogen and oxides of sulfur
in the ambient air.
       Following this introductory chapter, this document presents policy relevant information
drawn from the ISA and REA as well as assessments that translate this information into a basis
for staff conclusions as to policy options that are appropriate to consider in this review.  The
discussions are generally framed by addressing policy-relevant questions that have been adapted
from those initially presented in the IRP.
       Chapter 2 presents information that characterizes emissions, air quality, deposition and
water quality.  It includes discussions of the  sources of nitrogen and sulfur in the atmosphere as
well as current ambient air quality monitoring networks and models.  Additonal information in
this section includes ecological modeling and water quality data sources.
       Chapter 3 discusses the known or anticipated ecological  effects associated with oxides of
nitrogen and sulfur, including both deposition-related effects and well as direct effects. In so
doing, we address questions about the nature and magnitude of ecosystem responses to reactive
nitrogen and sulfur deposition, including responses related to acidification, nutrient enrichment,
and the  mobilization of toxic metals in sensitive aquatic and terrestrial ecosystems, and the
uncertainties and limitations associated with the evidence of such effects. Consideration is given
to how these responses are affected by landscape factors,  and what  types of ecosystems are
sensitive to such responses.  We also consider the extent to which ecosystem responses to
nitrogen deposition can be separated into responses related to oxidized and reduced forms of
reactive nitrogen compounds.
       In chapter 4, we address questions related to linking ecological effects to measures that
can be used to characterize the extent to which such effects are reasonably considered to be
adverse to public welfare. This involves consideration of how to characterize adversity from a
public welfare perspective. In so doing, consideration is given to the  concept of ecosystem
services, the evidence of effects on ecosystem services, and how ecosystem services  can be
linked to ecological indicators.
       Having focused more heavily on deposition-related effects on aquatic acidification in
chapters 3 and 4, chapter 5 considers the potential co-benefits that could be  expected to result
from a standard that is designed to provide protoection from aquatic acidification.  Consideration
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is given to potential co-benefits related to terrestrial acidification as well as aquatic and terrestrial
nutrient enrichment.
       Chapter 6 presents an assessment of the adequacy of the cuurent NC>2 and SC>2 seconday
standards. Consideration is given both to the adequacy of protection afforded by the current
standards for both direct and deposition-related effects, as well as to the appropriateness of the
fundamental structure and the basic elements of the current standards for providing protection
from deposition-related effects. In so doing, we address questions related to considering the
extent to which deposition-related effects that could reasonably be judged to be adverse to public
welfare are occurring under current conditions which are allowed by the current standards.  We
also consider the ways in which the structures and basic elements of the current NC>2 and SC>2
secondary standards  are inadequate to protect against such effects.
       Potential alternative standards for oxides of nitrogen and sulfur are considered in  chapter
7, drawing on the information in the previous chapters. More specifically, chapter 7 discusses
alternative approaches to defining the four elements of a NAAQS - indicator, form, averaging
time, and level - for  a standard intended to protect against effects on sensitive ecosystems
associated with deposition-related aquatic acification.  This chapter considers the implications of
alternative standards, including specific combinations of alternative  forms and levels, in terms of
identifying the sensitive ecosystems across the U.S. that would receive additional protection
from such alternative standards. Staff conclusions as to alterantive standards that are appropriate
to consider in this review are presented, together with the rationales  for such conclusions.
       This document also includes a number of appendices providing additional information to
support the document.  Appendix A provides an analysis conducted to compare aquatic
acidification to terrestrial acidification. Appendix B discuss critical  loads derivations and
modeling. Appendices C and D provide additional information regarding spatial aggregation and
critical loads. An overview of alternative indicators for oxides of nitrogen is located in
Appendix E and discussions of uncertainty analyses are included as Appendices F and G.
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1.6    REFERENCES

NAPAP.  1990. Acid Deposition: State of Science and Technology. National Acid Precipitation
       Assessment Program. Office of the Director, Washington, DC.
NRC (National Research Council). 2004. Air quality management in the United States.
       Washington, DC: National Research Council (NRC); The National Academies Press.
NSTC (National Science and Technology Council). 2005. National Acid Precipitation
       Assessment Program Report to Congress: An Integrated Assessment. Executive Office of
       the President, National Science and Technology Council, Washington, DC. Available at
       http ://www. esrl. noaa.gov/csd/aqrs/reports/napapreport05 .pdf
NSTC (National Science and Technology Council). 2010. National Acid Precipitation
       Assessment Program Report to Congress: An Integrated Assessment. Executive Office of
       the President, National Science and Technology Council, Washington, DC. Available at
       http ://www. esrl. noaa.gov/csd/aqrs/reports/napapreport 10.pdf
Russell, A and J. M. Samet. 2010a. Review of the Policy Assessment for the Review of the
       Secondary National Ambient Air Quality Standard for NOx and SOx: Second Draft.
       EPA-CASAC-11-003.
Russell, A and J. M. Samet. 2010b. Review of the Policy Assessment for the Review of the
       Secondary National Ambient Air Quality Standard for NOx and SOx: First Draft. EPA-
       CASAC-10-014.
U.S. EPA (Environmental Protection Agency).  1973. "Effects of Sulfur Oxide in the Atmosphere
       on Vegetation". Revised Chapter 5 of Air Quality  Criteria For Sulfur Oxides. U.S.
       Environmental Protection Agency. Research Triangle Park, N.C. EPA-R3-73-030.
U.S. EPA (Environmental Protection Agency).  1982. Review of the National Ambient Air
       QualityStandards for Sulfur Oxides: Assessment of Scientific and Technical Information.
       OAQPS Staff Paper. EPA-450/5-82-007. U.S. Environmental Protection Agency, Office
       of Air Quality Planning and Standards, Research Triangle Park, NC.
U.S. EPA (Environmental Protection Agency).  1984a. The Acidic Deposition Phenomenon and
       Its Effects: Critical Assessment Review Papers. Volume I Atmospheric Sciences. EPA-
       600/8-83-016AF. Office of Research and Development, Washington, DC
U.S. EPA (Environmental Protection Agency).  1984b. The Acidic Deposition Phenomenon and
       Its Effects: Critical Assessment Review Papers. Volume II Effects Sciences. EPA-600/8-
       83-016BF. Office of Research and Development, Washington, DC
U.S. EPA (Environmental Protection Agency).  1985. The Acidic Deposition Phenomenon and
       Its Effects: Critical Assessment Document. EPA-600/8-85/001. Office  of Research and
       Development, Washington, DC
U.S. EPA (Environmental Protection Agency).  1995a. Review of the National Ambient Air
       Quality Standards for Nitrogen Dioxide: Assessment of Scientific and Technical
       Information. OAQPS Staff Paper. EPA-452/R-95-005. U.S. Environmental Protection
       Agency, Office of Air Quality Planning and Standards, Research Triangle Park, NC.
       September.
U.S. EPA (Environmental Protection Agency).  1995. Acid Deposition Standard Feasibility Study
       Report to Congress. U.S. Environmental Protection Agency, Washington, DC.  EPA-
       430/R-95-001a.
U.S. EPA (Environmental Protection Agency).  2007. Integrated Review Plan for the Secondary
       National Ambient Air Quality Standards for Nitrogen Dioxide and Sulfur Dioxide. U.S.
       Environmental Protection Agency, Research Triangle Park, NC, EPA-452/R-08-006.

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U.S. EPA (Environmental Protection Agency). 2008. Integrated Science Assessment (ISA) for
       Oxides of Nitrogen and Sulfur Ecological Criteria (Final Report). U.S. Environmental
       Protection Agency, Washington, D.C., EPA/600/R-08/082F, 2008.
U.S. EPA (Environmental Protection Agency). 2009. Risk and Exposure Assessment for Review
       of the Secondary National Ambient Air Quality Standards for Oxides of Nitrogen and
       Oxides of Sulfur-Main Content - Final Report. U.S. Environmental Protection Agency,
       Research Triangle Park, NC, EPA-452/R-09-008a
Wolff, G. T. 1993. CAS AC closure letter for the 1993 Criteria Document for Oxides of Nitrogen
       addressed to U.S. EPA Administrator Carol M. Browner dated September 30, 1993.
Wolff, G. T. 1995. CASAC closure letter for the 1995 OAQPS Staff Paper addressed to U.S.
       EPA Administrator Carol M. Browner dated August 22, 1995.
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           2      CHARACTERIZING EMISSIONS, AIR QUALITY,
                     DEPOSITION AND WATER QUALITY

       This chapter provides an overview of air emissions, air quality, deposition, and water
quality relevant to oxides of nitrogen and sulfur, with specific focus on information related to
aquatic acidification processes (Figure 2-1).  Atmospheric, terrestrial and aquatic systems are
discussed, consistent with consideration of a multi-pollutant,  multi-media standard, which makes
the scope of this chapter much broader than to most NAAQS  policy assessments that
traditionally focus on atmospheric media only.  While serving as a general resource for data
availability and system descriptions (monitoring networks, models, emission inventories), the
information presented here provides background and context  for more focused policy-relevant
discussions in the subsequent chapters.  A source-to-effects continuum is adhered to in covering
the suite of topics, starting with emissions (section 2.1)  and proceeding through air quality
(section 2.2),  deposition (section 2.3), soils and  surface  waters (2.4), and followed by a summary
of trends for these four topics (section 2.5).
       Most of the atmospheric  and water quality based data presented here are intended to
reflect contemporary environmental conditions.  A 2005 base year is the most contemporary
atmospheric modeling available  and is used frequently in this assessment to characterize air
quality and deposition. While effort was made to  present recent water quality data, we note that
some of the most relevant water quality  data is of  late 1980s vintage.
                                Acidification of water + Eutrophication
Figure 2-1.  Overview of atmospheric, soil and aquatic processes relevant to acidification.
                                           2-1

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The multiple pollutant, multiple media technical systems framework
       The multiple pollutant, multiple media context of this assessment (Figure 2-2; Scheffe et
al, 2007, NARSTO, 2011) is based on the fact that air emissions move through the atmosphere
and are modified by chemical and physical reactions, advected and dispersed and ultimately
removed in the form of chemical deposition.   The multiple pollutant context not only is related
to the similarity of contributions to acidification from nitrogen and sulfur, but also to integration
of so many atmospheric species that influence nitrogen and sulfur patterns and, conversely, the
influences of nitrogen and sulfur on other air pollutant species of interest.  Consequently, there
are important linkages with other air pollutants and therefore other air management programs
and rules.   Because pollutants such as ozone and particulate matter are influenced by many of
the same emission sources and atmospheric processes, those programs are  of direct relevance,
from a chemical systems perspective, to assessments of oxides of nitrogen and sulfur.

                             Primary Sources
                                                             Chemical
                                                           «. Deposition
Figure 2-2 Diagram illustrating the multiple pollutant, multiple media linkages that are
incorporated in air quality models like CMAQ.  While this assessment focuses on the combined
effects of oxides of nitrogen and sulfur, the technical basis for characterizing spatial and
temporal patterns of N and S is dependent on several other atmospheric species.
                                           2-2

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Terminology, definitions and units.
       Throughout this document numerous terms are used that address a variety of atmospheric
and ecosystem processes and variables.  We establish the terminology here, early in the
document, as a reference source for the entire report.
       As discussed in detail in the REA (REA 1.3.1), in the atmospheric science community
NOx is typically referred to as the sum of nitrogen dioxide (NCh), and nitric oxide (NO).  The
term used by the scientific community to represent the complete set of reactive oxidized nitrogen
compounds is total oxidized nitrogen (NOy), commonly defined as NO, NO2 and the all of the
oxidation products of NO and NO2 Reactive oxidized nitrogen is defined as NOy = NO2 + NO
+ HNO3 + PAN +2N2O5 + HONO+ NO3 + organic nitrates + particulate NO3 (Finlayson-Pitts
and Pitts, 2000).  In this document, unless otherwise indicated, we use the term NOy as the
atmospheric indicators associated with the NOx component of the proposed NOx/SOx standard.
       For this assessment, SOx is defined to include all oxides of sulfur, including multiple
gaseous substances (e.g., SO2, sulfur monoxide [SO], sulfur trioxide [SO3], thiosulfate [S2O3],
and heptoxide [820?],  as well as particulate species, such as ammonium sulfate [(NH4)2SO4]).
Throughout this text we refer to sulfate as SO4 and nitrate as NO3, recognizing that they have
charges of-2 for sulfate and -1  for nitrate. The sum of sulfur dioxide gas (802) and particulate
sulfate (804), referred herein as (SO2 + 804) is used throughout this document as the
atmospheric indictor for the SOx component of the proposed NOx/SOx standard. From a
measurement and  modeling perspective we only consider the sum of SO2 and particulate SO4 as
the indicator for sulfur.  The sum of SO2 and SO4 constitute virtually  all of the ambient air sulfur
budget and are measured routinely in monitoring networks.
       Table 2-1  provides further explanation of these indicators, some of which is repeated in
Section 7.  Table  2-1 also provides details on the units used throughout the equations and
examples in the PA.   Again, because of difference in unit conventions between atmospheric and
ecosystem sciences, there are detailed explanations of units as well as  procedures for translating
between different  unit conventions. To facilitate the  linkage between atmospheric and
ecosystem processes, only the mass (or equivalent charge)  associated with sulfur or nitrogen is
considered in mass, mixing ratio,  and  deposition unit conventions.

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Table 2 -1. Description of parameters, units and conventions.
Parameter
Units Conversions to other unit
conventions used in figures and
calculations (multiply value in Units
column by:
Explanation
Atmospheric species
CMAQ defined NOy species: NO (nitrogen oxide), NO2 (nitrogen dioxide), HNO3 (nitric acid), p-NO3 (participate bound nitrate), NO3 (sum of
HNO3 and p-NO3), PAN (peroxy acetyl nitrate), N2O5 (dinitrogen pentoxide), PANX (higher order PANs), NTR (organic nitrates), PNA (HNO4);
sulfur dioxide (SO2), particulate sulfate (SO4);
NHX species: NH3 (ammonia), ammonium ion (NH4)
Lumped Atmospheric Species
NO,
(S02 +S04)
NHX
The sum of all reactive oxidized nitrogen compounds derived through summing all nitrogen
contributions (i.e., 1-HNO3 + 2-N205 +. . .) from the modeled species (HNO3, p-NO3, NO2, NO, PAN . . .)
or through direct measurement which reduces all oxidized nitrogen species to NO and reports as ppb NO.
All references to the quantity NOy refer to the mass, molar or equivalent charge contribution of
nitrogen only. All mass contributions of oxygen, hydrogen and carbon are not included.
Oxidized forms of sulfur defined as sulfate (SO4 + SO2); mass units maintained for consistency with
deposition calculations
Note that only mass as sulfur is counted in state variables; in practice, individual SO2 and SO 4 are
measured/modeled and converted to mass of sulfur atoms or equivalent charge units. Mass
contribution of oxygen is not included.
Reduced nitrogen calculated as the sum of NH3 and NH4. All references to the quantity NHX used as
state variables refer to the mass, molar or equivalent charge contribution of nitrogen only. Mass
contribution of oxygen is not included.
Atmospheric State Variables used in equations and derivations
NOS concentration
sox
NHX
Used in various conventions of:
C,;
NOy deposition
(S02 +S04)
NHX
Used in various conventions of:
Depi
Ndep
NOydep
NHx
Sdep
Wet
v .
i
Dep^
fig/m as N or S
meq/m2 -yr as N or S
meq/m2 -yr
m/yr
tneq/nf-yr
ppb = (MA/Mj)- pair) -fig/m
where pajr is the air density in units of
(kg/m3);
Pair = 28.97( 10)'3-P/(R-T)
R= 8.206(10)'5m3atm/(mol-K)
P = atm
T = degrees K
MA = molecular weight of air (28.97)
MJ= Atomic weight of nitrogen (14) or
sulfur (32)
meq/m3 = (1/Mj) -fig/m3
kg/ha-yr =(Mj/q)(10)" • meq/m -yr
where q = charge (1 for N, 2 for S)





Total (wet and dry) deposition;
Ndep = NHx + NOYdep
Sdep = SOx=SO2 +SO4
wet deposition velocities
dry deposition fluxes
                                     2-4

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Parameter
Wet
Dep.

DepJ°tal
NOy deposition
(S02 +S04)
NH,
Used in various conventions of:
Depi
Ndep
NOydep
NHx
Sdep
vF
vr
Depp1
Depr
Dep,total
TSO*
TNOV
Units Conversions to other unit
conventions used in figures and
calculations (multiply value in Units
column by:
meq/m -yr

meq/m -yr
meq/m2 -yr as N
orS
meq/m2 -yr
m/yr
m/yr
meq/m -yr
meq/m -yr
meq/m -yr
m/yr



Ag^a-jr=(M1/q)(10)"2 • meq/m2-yr
where q = charge (1 for N, 2 for S)






Calculated by dividing total
(SO2 +SO4) or NOy deposition
(wet and dry) by the annual
average (SO2 +SO4) or NOy
concentration.
Explanation
wet deposition fluxes

total (wet+dry) deposition

Total (wet and dry) deposition;
Ndep = NHx + NOYdep
Sdep = SOx=SO2 +SO4
dry deposition velocities
wet deposition velocities
dry deposition fluxes
wet deposition fluxes
total (wet+dry) deposition
the transfer ratio, which can be
considered an aggregated,
"effective" deposition velocity
that relates total deposition of
(SO2 +SO4) or NOy to the total
ambient concentration, and
represents an average of the
chemical species specific v™ ( =
vDry + vW,t^ values
Ecosystem variables
ANC
ANCimit
*-'-L'anclim(,lJ
CL(N+S)
CL(S)
Q
Qr
NECO
Nleach
CLr
[leq/L
[leq/L
meq/m2-yr
m/yr
m/yr
meq/m -yr
meq/m2-yr
meq/m -yr








measured ANC in surface water
a "target" ANC level
Critical load that does not cause
the catchment to exceed a given
ANCiim- where ;' indicates the
pollutant of interest
Average surface water runoff rate
for a water body
Median of the average runoff rates
for water bodies in an ecoregion
Nitrogen uptake, retention and
denitrification by terrestrial
catchment
N leaching based on observed
surface water NO3
Ecoregion representative critical
load
2-5

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2.1    SOURCES OF NITROGEN AND SULFUR
       This section recasts much of the information provided in EPA's Risk and Exposure
Assessment (EPA, 2009). The emission summaries are based on the 2002 calendar year and are
intended to convey the basic patterns and major contributors of NOx, SOx and NH3 emissions.
The air quality modeling simulations used in chapter 7, as well as some of the air quality and
deposition illustrations in this chapter are based on a more modern 2005 calendar year
simulation.   For the purposes of presenting general patterns of emissions, the 2002 emissions
presented here are  not significantly different than the 2005 year data.
        The National Emissions Inventory (NEI) annual total emissions data for 2002 (U.S.
EPA, 2006) are used to characterize the magnitude and spatial patterns in emissions of NOX,
NH3, and 862 nationwide1. The spatial  resolution of these data varies by source type. Emissions
from most large stationary sources are represented by individual point sources (e.g., electric
generating units, industrial boilers).  Sources that emit over broad areas are reported as county
total emissions. The national annual 2002 emissions of NOX, NHa, and SC>2 by major source
category are presented in Table 2-1 of the ISA (U.S. EPA, 2008).
2.1.1  NOX Emissions
       The distribution of national total NOX emissions across major source categories is
provided in Table 2-2.  Emissions summaries are also provided for the East2 and West in Tables
2-3 a and b, respectively, to reveal regional differences in source emissions profiles. In addition
to anthropogenic sources, there are also natural sources of NOX, including lightning, wildfires,
and microbial activity in  soils. Nationally, transportation-related sources (i.e., on-road, nonroad,
and aircraft/locomotive/marine) account for -60% of total anthropogenic emissions of NOX,
while stationary sources (e.g., electrical utilities and industrial boilers) account for most of the
remainder (U.S. EPA, 2008, AX2, Table 2-1). Emissions from on-road vehicles represent the
major component of mobile source NOX emissions. Approximately half the mobile source
emissions are contributed by diesel engines, and half are emitted by gasoline-fueled vehicles and
other sources (U.S. EPA, 2008, AX2, Section 2.1.1 and Table 2-1-1). Nationwide,  the nonroad,
1 For the purposes of this analysis, nationwide emissions do not include emissions from Alaska or Hawaii.
2 In this analysis, the East is defined as all states from Texas northward to North Dakota and eastward to the East
Coast of the United States. States from New Mexico northward to Montana and westward to the West Coast are
considered to be part of the West.
                                            2-6

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aircraft/locomotive/marine, and non-electric generating unit point emissions sectors each
contribute generally similar amounts to the overall NOX inventory. Overall, NOX emissions are
broadly split between NO and NO2 in a ratio of 90% NO and  10% directly emitted NO2.
However, this split can vary by source category, as described  in Chapter 2.2.1 of the ISA (U.S.
EPA, 2008).

Table 2-2. Annual National NOX Emissions across Major Source Categories in 2002.
National Totals
Electric Generation Units
Industrial Point Sources
Stationary Area
On-road
Nonroad
Aircraft/Locomotive/Marine
Fires
Total
NOX
Emissions (million tons)
4.619
2.362
1.529
7.839
2.219
2.611
0.080
21.259
Percent of Total
22%
11%
7%
37%
10%
12%
< 1%

Table 2-3a. Annual NOX Emissions across Major Source Categories in 2002 for the Eastern
United States.
Eastern U.S.
Electric Generation Units
Industrial Point Sources
Stationary Area
On-road
Nonroad
Aircraft/Locomotive/Marine
Fires
Total
NOX
Emissions (million tons)
4.094
2.031
1.295
6.250
1.709
2.038
0.028
17.445
Percent of Total
23%
12%
7%
36%
10%
12%
< 1%

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Table 2-3b. Annual NOX Emissions across Major Source Categories in 2002 for the Western
United States.
Western U.S.
Electric Generation Units
Industrial Point Sources
Stationary Area
Onroad
Nonroad
Aircraft/Locomotive/Marine
Fires
Total
NOX
Emissions (million tons)
0.525
0.331
0.234
1.589
0.510
0.573
0.055
3.817
Percent of Total
14%
9%
6%
42%
13%
15%
1%

       In general, NOX emissions in the East are nearly 5 times greater that NOX emissions in the
West. In both the eastern and western United States, the on-road sector is the largest contributor.
Emissions from electric generation units are the second-largest contributor to NOX emissions in
the East with 23% of the total. Emissions in the East from industrial point sources, nonroad
engines,  and aircraft-locomotives-marine engines each contribute in the range of 10 to 12%. In
the West, the contribution to NOX emissions from electric generation units (14%) is in the same
range as  the contributions from nonroad engines (13%) and aircraft-locomotives-marine engines
(15%).
       The  spatial patterns of 2002 annual NOX emissions across the United States are shown in
Figure 2-33. Emissions of NOX are concentrated in and near urban and suburban areas and along
major highways. Moderate or higher levels of NOX emissions (>100,000 tons/yr)4 are also
evident in some rural areas at locations (i.e., grid cells) containing major point sources. The
amount of NOX emissions in and near each of the case study areas can be seen from this map. All
of the case study areas contain or are near locations with NOX emissions in excess of
100,000 tons/yr.
3 To create this map, NOX emissions were allocated to a 36 x 36- km grid covering the United States in order to
normalize for the differences in the geographic aggregation of point- and county-based emissions. The emissions are
in tons per year per 36 x 36 km (1,296 km2).
4 Emissions are in tons per year per 36 x 36 km (1,296 km2).
                                            2-8

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      Legend
      NOx tons/year
      white =0.0
         > Q to 25,000
         > 25,000 to 100.000
         > 100,000 to 250,000
      ^| > 250.000 to 500.000
      ^B > 500,000 to 1000,000
      ^H >1.000.000 to 2,456,200
. •-•
1 Adirondack
2 Shenandoah
3 Potomac River/Potomac Estuary
4 Neuse River/Neuse Estuary
5 Kane Experimental Forest
6 Hubbard Brook Experimental Forest
7 Mixed Conifer Forest (Transverse Range)
8 Mixed Conifer Forest (Sierra Nevada Range)
9 Rocky Mountain National Park
        Figure 2-3. Spatial distribution of annual total NOX emissions (tons/yr) for 2002.
2.1.2   NH3 Emissions
        The primary anthropogenic sources of NH3 emissions are fertilized soils and livestock.
Motor vehicles and stationary combustion are small emitters of NIT?. Some NIT? is emitted as a
byproduct of NOX reduction in motor vehicle catalysts. The spatial patterns of 2002 annual NHs
emissions are shown in Figure 2-45. The highest emissions of NtT? are generally found in areas
of major livestock feeding and production facilities, many of which are in rural areas.  In
addition, MT? emissions exceeding 1,000 tons/yr are evident across broad areas that are likely
associated with the application of fertilizer to crops. The patterns in NHs  emissions are in
5 Note that, because overall emissions of NH3 are much lower than emissions of NOX, we used a more refined set of
ranges to display emissions of NH3 compared to what was used to display emissions of NOX.
                                                2-9

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contrast to the more urban-focused emissions of NOX. The Potomac River/Potomac Estuary,
Neuse River/Neuse River Estuary,  Shenandoah, and Mixed Conifer Forest (in the Sierra Nevada
Range and the Transverse Range) case study areas all have sources with NHs emissions
exceeding 5,000 tons/yr. Rocky Mountain National Park is adjacent to an area with relatively
high NHa emissions exceeding 2,500 tons/yr. The Adirondack, Hubbard Brook Experimental
Forest, and Kane Experimental Forest case study areas are more distant from sources of NH3 of
this magnitude.
     Legend
     NH3 tons/year
     white =00
         > o to 100
         > too to 1.000
         > 1.000 to 2,500
     IB > 2.500 to 5.000
     ^| > 5.000 to 10,000
     •• > 10,000 to 21.908
1 Adirondack
2 Shenandoah
3 Potomac River/Potomac Estuary
4 Neuse River/Neuse Estuary
5 Kane Experimental Forest
6 Hubbard Brook Experimental Forest
7 Mixed Conifer Forest (Transverse Range)
8 Mixed Conifer Forest(Sierra Nevada Range)
9 Rocky Mountain National Park
       Figure 2-4. Spatial distribution of annual total NHs emissions (tons/yr) for 2002.
2.1.3  SOX Emissions
       The distributions of SC>2 emissions for major source categories nationally are provided in
Table 2-4. Emissions of 862 for the East and West are presented in Tables 2-5 a and b,
respectively. Anthropogenic  emissions of 862 in the United States are mainly due to combustion
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of fossil fuels by electrical generation units (70%) and industrial point sources (15%);
transportation-related sources contribute minimally (7%). Thus, most SC>2 emissions originate
from point sources. Almost all the sulfur in fuel is released as volatile components (SC>2 or SOs)
during combustion. The higher sulfur content of coal compared to other types of fossil fuels
results in higher 862 emissions from electrical utilities using coal as fuel.
       Emissions of SC>2 are more than 10 times greater in the East than in the West. Emissions
from electric generation units are the largest contributor to 862 emissions in both the East and
West, but are a much greater fraction of the inventory in the East (71%) compared to the West
(44%). Stationary area sources and the aircraft-locomotive-marine engine sector have a greater
relative contribution  to SC>2 in the West compared to the East6.
       The largest natural sources of SC>2 are volcanoes and wildfires. Although SC>2 constitutes
a relatively minor fraction (0.005% by volume) of total volcanic emissions (Holland,  1978),
concentrations in volcanic plumes can be range up to tens of parts per million (ppm). Sulfur is a
component of amino acids in vegetation and is released during combustion. Emissions of 862
from burning vegetation are generally in the range of 1% to 2% of the biomass burned (Levine et
al,  1999).
 Note that SO2 emissions from fires are understated in the NEI because of an error in the emissions calculations.
                                           2-11

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Table 2-4. Annual National SO2 Emissions across Major Source Categories in 2002.
National Totals
Electric Generation Units
Industrial Point Sources
Stationary Area
On-road
Nonroad
Aircraft/Locomotive/Marine
Fires
Total
SO2
Emissions (million tons)
10.359
2.249
1.250
0.242
0.188
0.533
0.050
14.871
Percent of Total
70%
15%
8%
2%
1%
4%
<1%

Table 2-5a. Annual SC>2 Emissions across Major Source Categories in 2002 for the Eastern
United States.
Eastern U.S.
Electric Generation Units
Industrial Point Sources
Stationary Area
On-road
Nonroad
Aircraft/Locomotive/Marine
Fires
Total
SO2
Emissions (million tons)
9.923
2.057
1.116
0.214
0.162
0.398
0.011
13.881
Percent of Total
71%
15%
8%
2%
1%
3%
< 1%

                                         2-12

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Table 2-5b. Annual SC>2 Emissions across Major Source Categories in 2002 for the Western
United States.
Western U.S.
Electric Generation Units
Industrial Point Sources
Stationary Area
On-road
Nonroad
Aircraft/Locomotive/Marine
Fires
Total
SO2
Emissions (million tons)
0.436
0.192
0.134
0.029
0.026
0.136
0.035
0.988
Percent of Total
44%
19%
14%
3%
3%
14%
4%

       The spatial patterns of 2002 annual SC>2 emissions are shown in Figure 2-5. High SC>2
emissions are scattered across the East, and there are large sources in both urban are rural
locations. The greatest geographic concentration of 862 sources is in the Midwest, particularly
along the Ohio River, where numerous electric generating units are located. As noted above, 862
emissions in the West are much lower than in the East, with sources concentrated in urban
locations along with localized emissions in more rural areas associated with industrial sources
(e.g.,  smelters) and gas-field operations.
       The Potomac River/Potomac Estuary, Neuse River/Neuse River Estuary, Shenandoah,
and Mixed Conifer Forest (Transverse Range portion) case study areas each contain numerous
locations of major 862 emitters.  The Kane Experimental Forest Case Study Area and Rocky
Mountain National Park are relatively close to SOX emission locations exceeding 5,000 tons/yr.
The Adirondack, Hubbard Brook Experimental Forest, and Mixed Conifer Forest (Sierra Nevada
Range portion) case study areas are more distant from SOX sources of this magnitude.
                                          2-13

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     Legend
     S02 tons/year
     white =00
         >0to100
         > 100 to 1.000
         > 1.000 to S.000
         > 5.000 to 25,000
     ^| > 25.000 to 100.000
     ^H > 100 000 to 245.740
1 Adirondack
2 Shenandoah
3 Potomac River/Potomac Estuary
4 Neuse River/Neuse Estuary
5 Kane Experimental Forest
6 Hubbard Brook Experimental Forest
7 Mixed Conifer Forest (Transverse Range)
8 Mixed Conifer Forest (Sierra Nevada Range)
9 Rocky Mountain National Park
        Figure 2-5. Spatial distribution of annual total SC>2 emissions (tons/yr) for 2002.

2.2     AMBIENT AIR CHARACTERIZATION
        Characterizing air quality that is relevant to a secondary standard for oxides of nitrogen
and sulfur in which effects are transmitted from the air to aquatic systems through deposition
should include the species related to oxides of nitrogen and sulfur that have the potential to
contribute to aquatic acidification.   Because many of the atmospheric species that contribute to
acidification are not measured routinely or the monitoring networks have relatively  sparse spatial
coverage, we take advantage of CMAQ to illustrate patterns of different atmospheric species.
For the same reason, we also use CMAQ  to characterize deposition, particularly dry deposition.
               At certain points in this section, we describe certain basic concepts of air quality,
deposition, soils and water chemistry processes to provide the background prior for developing
the form of the standard in chapter 7.  For example, we start here by defining the relevant
species of NOy and explain why NOy is relevant to acidification by introducing explaining the
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relationship of atmospheric species to equivalent charge, which is fundamental to linking
atmospheric and aquatic systems.  Also, we separate air quality and deposition in order to
provide a more focused discussion on deposition processes and concepts which incorporated in
engineering the form of the standard as described in chapter 7.
       The key pollutants for this assessment are total oxidized nitrogen (NOy), total reduced
nitrogen (NHx), and total oxidized sulfur which is referenced herein as (SOx) and defined as the
sum of SO2 (gas) and particulate sulfate, as described above. Total reactive oxidized
atmospheric nitrogen, NOy, is defined as NOx (NO and NO2) and all oxidized NOx products:
NOy = NO2 + NO + HNO3 + PAN +2N2O5 + HONO+ NO3 + organic nitrates + particulate NO3
(Finlayson-Pitts and Pitts, 2000).  This definition of NOy reflects the operational principles of
standard measurement techniques in which all oxidized nitrogen species are converted to
nitrogen oxide (NO) through catalytic reduction and the resulting NO is detected through
luminescence.  Thus, NOy is truly defined as total oxidized nitrogen as converted to NO,
essentially  representing all oxidized nitrogen atoms. NOy is not a strict representation of the all
moles of oxidized nitrogen as the diatomic nitrogen species such as N2Os yield 2 moles of NO.
This definition is consistent with the relationship between atmospheric nitrogen and acidification
processes as the reported NOy provides a direct estimate of the potential equivalents available
for acidification.  We emphasize NOy here as all of the individual NOy species are potential
contributors to acidic deposition.  All NOy species are derived directly from NOx emissions or
through atmospheric transformations, thus establishing a direct link to oxides of nitrogen as they
are considered listed pollutants in the CAA.
       Total reduced nitrogen (NHx) includes ammonia, NH3, plus ammonium, NH4 (EPA,
2008) is introduced because NHx contributes potentially acidifying deposition, effectively
behaving similarly to NOy.  While NOy is not treated the same way as NOy in developing the
form,  it is incorporated because NHx can contribute to acidifying deposition.  Reduced nitrogen
plus oxidized nitrogen is referred to as total reactive nitrogen.  Total oxidized sulfur (SOx)
includes SO2 gas and particulate  sulfate, SO4. These species are converted to mass of sulfur
which is used directly, or converted to charge equivalents, in deposition analyses linking
atmospheric deposition and ecosystem models.  Ammonium and sulfate are components of
atmospheric particulate matter as well as directly measured and modeled in precipitation as
direct deposition components.
                                           2-15

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       The term charge equivalents refers to positively charged cations (e.g., Mg+2, Ca+2, H+, K+,
Na+) or negatively charged anions (NO3-, SO4-2, C1-, OH-) in solution.  Any nitrogen atom in the
NOy or NHx species mix has the potential to provide one negative charge. Effectively, this
means that regardless of the specific nitrogen compound structure, the nitrogen atom eventually
can be transformed into nitrate, NO3- that enters an aquatic system. Similarly, for sulfur any
atmospheric sulfur atom has the potential to be transformed into a sulfate ion, SO4"2, which has
an equivalent charge of-2.  For convenience,  we use NOs and SO4 to represent nitrate and
sulfate ion, respectively.  One can consider that each atom of sulfur offers twice the acidifying
potential of any nitrogen atom. In this sense of recognizing the unique importance of nitrogen
and sulfur atoms, there is  a direct connection with emissions of oxides of nitrogen or sulfur as
described by the CAA. In other words, every atom of emitted nitrogen in  NOx emissions
remains in the atmosphere as a component of NOy, irrespective of whether the species attached
to a nitrogen atom is the same as emitted originally, or transformed to another form in the
atmosphere.
       Further discussion of the processes in the atmosphere and terrestrial and aquatic systems
responsible for the transformations of nitrogen and sulfur species to NOs and SO4 are briefly
discussed in this section and in appendices where noted.
2.2.1   Air monitoring networks
       There are over 1000 ground level monitoring platforms (Figures 2-6 and 2-7 and Table 2-
6) that provide measurements of some form of atmospheric nitrogen or sulfur.
                                          2-16

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Table 2-6. Summary of Monitoring Networks.
Network
Number
of Sites
Species Measured
Sampling
Frequenc
y
Comments
All Sulfur Sites
NCore
SEARCH
SO2
PM Speciation
IMPROVE
CASTNET
82
8
751
242
215
88
SO2
SO2
SO2
Sulfates
Sulfates
Sulfates
Hourly
Hourly
Hourly
24-hour
24-hour
Weekly
Ave.
Includes 20 rural sites
Includes 3 rural sites
NAMS/SLAMS/PAMS
for 2008
Measurements of
Sulfates (88403)
identified in AQS for
Trends and
Supplemental Speciation
monitoring type for
2008
IMPROVE Monitoring
Sites with Measurements
of Sulfates (88403)
identified in AQS
EPA & NPS
All Nitrogen Sites
NCore
SEARCH
PAMS
SLAMS
NOY
IMPROVE
CASTNET
AMON
82
8
119
643
59
214
88
-20
NO/NOV
NO/NO2/NOV/HNO3
NO2/NOx
NO/NO2/NOx/NOy
NOy
Nitrates
Nitrates
NH3
Hourly
Hourly
Hourly
Hourly
Hourly
24-hour
Weekly
average
Monthly
average
Includes 20 rural sites
Includes 3 rural sites
Official sites as of 12/09
All SLAMS Monitoring
Sites with Measurements
of NO, NO2, NOX or
NOY in 2009 identified
in AQS
All Monitoring Sites
with Measurements of
NOY in 2009 identified
in AQS, regardless of
Monitoring Type
MPROVE Monitoring
Sites with Measurements
of Nitrates (88306)
identified in AQS
EPA & NPS
New program
component of NADP;
passive sampling
technique
                                     2-17

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                                       MAPI (All N)
            NCore, NOY(2009), SEARCH, PAMS/SLAMS, CASTNET, IMPROVE
      Legend
        NOY_2009
        NCore
        Rural NCore
        SEARCH
        Rural SEARC
        PAMS_NQ-N02-NOX-NOY_2009
        SLAMSJJ 0-N02-N QX-NQY
        CASTNET-NPS
        CASTNET-EPA
        IMPROVE Nitrates 2006
Figure 2-6.   Routinely operating surface monitoring stations measuring forms of
                atmospheric nitrogen.
                                               2-18

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                                     MAP6-1a(AII S)
                 NCore, SO2(2008), SEARCH, CASTNET, IMPROVE, and
                      Trends/Supplemental Speciation Sites (2008)
                        -- SO2(2008) includes NAMS / SLAMS / PAMS --
       Legend
       •  S02_2008
      -*-  NCore
      O  Rural NCore
       •  SEARCH
      D  Rural SEARCH
       •  CASTNET-NPS
         CASTNET-EPA
       A  Speciation_Sulfates_2008
       »  IMPROVE _Sulfstes_2006
Figure 2-7.   Routinely operating surface monitoring stations measuring forms of atmospheric
              sulfur.  All site locations measure both SC>2 and sulfate except for the green SC>2
              only sites.
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       As discussed in this section, there are only very limited routine measurements of total
oxidized and reduced nitrogen.  In addition, existing monitoring networks do not provide
adequate geographic coverage to fully assess concentrations and deposition of reactive nitrogen
and sulfur in and near sensitive ecosystems.
       The principal  monitoring networks include the regulatory based State and Local Air
Monitoring Stations (SLAMS) providing mostly urban-based 862, NO and NOx, the PM2.5
chemical speciation networks Interagency Monitoring of Protected Visual Environments
(IMPROVE) and EPA's Chemical Speciation Network (CSN) providing particle bound sulfate
and nitrate, and the Clean Air Status and Trends Network (CASTNET) providing weekly
averaged values of SO2, nitric acid, and particle bound sulfate, nitrate and ammonium.  The
private sector supported South Eastern Aerosol Research and Characterization (SEARCH) Study
network of 4-8 sites in the Southeast provides the only routinely operating  source of true
continuous NO2, ammonia, and nitric acid measurements. SEARCH also provides PM2.5 size
fractions of nitrate and sulfate.  Collectively, the SLAMS, Photochemical Assessment
Measurement Stations (PAMS), SEARCH and NCore networks will provide over 100 sites
measuring NOy (Figure 2-8).  The NCore network (Scheffe et al, 2009) is  a multiple pollutant
network with co-located measurements of key trace gases (CO, SO2, Os, NO and NOy), PM2.5
and PM (io-2.5) mass and PM2.s chemical speciation.  Additional air pollutants, particularly volatile
organic compounds (VOCs), will be measured at those sites that are part of the existing PAMS
and National Air Toxics Trends (NATTS) platforms.  The NATTS (EPA, 2008) include 27
stations across the U.S. that monitor for  a variety of hazardous air pollutants and are intended to
remain in place to provide a long-term record.  Additional measurements of ammonia and
possibly true NO2 are under consideration.  True NO2 is noted to differentiate from the NO2
determined through routine regulatory networks that have known variable positive bias for NO2.
       The network currently is being deployed and expected to be operational with nearly 75
sites by January 2011.  The sites are intended to serve as central site monitors capturing broadly
representative (e.g., not strongly influenced by nearby sources) air quality in a suite of major and
mid size cities and approximately 20 sites are located in rural locations.
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                                        MAP 4
                   Current & Planned Routine NOY Monitoring Sites
                             NCore, NOY(2009), SEARCH
Figure 2-8.   Anticipated network of surface based NOy stations based on 2009 network design
             plans.  The NCore stations are scheduled to be operating by January, 2011.

       There are significant measurement gaps for characterizing NOy, NHX and 862 in the
nations ambient air observation networks (EPA, 2008) that lead to greater reliance on air quality
modeling simulations to describe current conditions. National design of routinely operating
ambient air monitoring networks is driven mostly by data uses associated with implementing
primary NAAQS, with noted exceptions of the CASTNET and IMPROVE networks.  In
addition to significant spatial gaps in sensitive ecosystem areas that arise from a population
oriented network design, the current measurements for primary and secondary nitrogen are
markedly different and in some instances of negligible value for secondary NOx and SOx
standards. For example, a true NOx (NO plus NO2) measurement typically would capture less
than 50% (see discussion below) of the total regional NOy mass in rural locations as the more
aged air masses contain significant oxidized nitrogen products in addition to NOx.  With the
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exception of the SEARCH network in the Southeast, there have been virtually no routine
networks that measure ammonia until the recent addition of the AMON passive ammonia
sampling network (Figure 2-9) as part of the NADP. EPA is considering adopting the AMON
passive sampling techniques and other ammonia sampling options in the NCORE network.
Ammonium is reported in EPA chemical speciation networks, although the values are believed to
be biased low due to ammonia volatization.
       CASTNET provides mostly rural measurements of 862, total nitrate, and ammonium,  and
affords an existing infrastructure useful for future monitoring in support of a potential NOx and
SOx secondary standard. However, the lack of NOy, SOx and NHx measurements in sensitive
ecosystems will require attention in conjunction with any rulemaking for a secondary standard
for oxides of nitrogen and sulfur.
       As a result of the limited monitoring networks forNOy and SOx in sensitive ecosystems,
we are unable to use current ambient monitoring data to adequately link measured current
atmospheric concentrations to ecological effects transmitted through deposition. At this time for
the purpose of illustrating current atmospheric conditions, we supplement the available
monitoring  data with the use of sophisticated atmospheric modeling conducted using EPA's
CMAQ model (as discussed in chapter 7).
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             AMolT Site  Map and Modeled NH3  Emissions
  Modeled NH3 Emissions (kg/
                                     "A" Active Sites
                                     •fr Inactive Sites
                                                          Source US EPA and NADP
Figure 2-9.  Location of NADP passive ammonia sampling locations for the AMON network
(http: //nadp. sws. uiuc. edu/nh3 net/).
       Of the currently operating monitoring networks, precipitation based sulfate, ammonium
and nitrate measurements provided by the NADP are the most relevant measurements that would
support the secondary standard as they provide atmospheric deposition inputs that drive
ecosystem models, and NADP site locations generally include acid sensitive areas.  However,
there are significant gaps in ambient air (aerosols and gases) monitoring networks for the
measurement of the likely ambient indicators of NOy, 862, and 864. CASTNET filter packs
provide the most relevant source of ambient sulfate (804) measurements as the open inlet of the
filter packs incorporates the full range of particle sizes that contribute to deposition.  The SC>2
measurements from CASTNET represent about 10% of all SC>2 sites nationally, but are
especially relevant based on their locations  in rural and regional settings, although CASTNET  is
not as spatially extensive (breadth and resolution) as the NADP network of precipitation sites.
Although CASTNET does provide measurements of total ambient nitrate, other oxidized
nitrogen species constituting a more complete NOy budget are not captured. In their current
configuration, the State and local monitoring networks offer virtually no support for a secondary
                                          2-23

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standard for oxides of nitrogen and sulfur due to their urban-based site orientation and exclusion
of important oxidized nitrogen species (e.g., nitrates and PAN).  The chemical speciation
networks, including rural based IMPROVE, all provide ambient sulfate measurements based on a
2.5|i size cut. While the sulfate mass within that size fraction may constitute 80% or greater of
the ambient sulfate budget, the missing larger size particles can contribute significantly to sulfate
deposition due to their relatively high gravitationally driven deposition velocities. Finally, there
are virtually no ambient ammonia measurements routinely collected in acid sensitive areas.
CASTNET does provide ammonium measurements, but the routine speciation networks that
report ammonium have expected artifacts due to ammonia off-gassing from nylon filters.
       Although this summary of existing networks suggests significant challenges in meeting
the monitoring needs of a new standard for oxides of nitrogen and sulfur, the networks do serve
as a useful building block for moving forward.  The site locations of NADP and CASTNET offer
an infrastructure to accommodate additional instruments. The NCORE network has introduced
nearly 75 NOy trace level 862 monitors that are establishing operational familiarity and a basis
for instrument performance characterization.  In many cases, acid sensitive areas will be strongly
influenced by regional transport of pollutants which typically is associated with relatively
homogeneous spatial concentration patterns which allows for a correspondingly greater range of
spatial representativeness of monitoring sites.  Consequently, the expected burden on monitoring
resources  may be realistically dampened by the available infrastructure and expected
homogeneity of air concentration patterns.  A more thorough assessment of the adequacy of
existing networks is predicated on identification of the area wide boundaries of the acid sensitive
areas of concern which will initially e developed in the second PAD.

2.2.2  Overview of CMAQ
       The Community Multiscale Air Quality (CMAQ) model was used to characterize air
quality and deposition.  CMAQ simulates the numerous physical and  chemical processes
involved in the formation, transport, and destruction of ozone, particulate matter and air toxics.
In addition to the CMAQ model, the modeling platform includes the emissions, meteorology,
and initial and boundary condition data which are inputs to this model.
       The 2005-based CMAQ modeling platform was used as the basis for national maps of air
quality and deposition reflect 2005 year meteorology and emissions.  An emissions sensitivity
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simulation with domain reductions of 48 and 42 % for NOx and SOx, respectively, was used to
explore the behavior of the form of the standard to potential future changes in air quality
associated with potential changes in emissions, and those results are discussed in chapter 7 and
Appendix E.  This platform represents a structured system of connected modeling-related tools
and data that provide a consistent and transparent basis for assessing the air quality response to
projected changes in emissions.  The platform was developed by the EPA's Office of Air Quality
Planning and Standards in collaboration with the Office of Research and Development and is
intended to support a variety of regulatory and research model applications and analyses.
       The CMAQ model is a comprehensive, peer-reviewed, three-dimensional grid-based
Eulerian air quality model designed to simulate the formation and fate of gaseous and particle
(i.e., particulate matter or PM) species, including ozone, oxidant precursors, and primary and
secondary PM concentrations and sulfur and nitrogen deposition over urban, regional, and larger
spatial scales ( Byun and Schere, 2006). CMAQ is run for user-defined input sets of
meteorological conditions and emissions.
       Additional details of the modeling domain, emissions and meteorological inputs are
provided in EPA (2009; REA Appendices).
Model domain and grid resolution.
       CMAQ modeling analyses were performed for a domain covering the continental United
States, as shown in Figure 2-10 and Table 2-7. This domain has a parent horizontal grid of 36
km with two finer-scale 12 km grids over portions of the eastern and western U.S.  The model
extends vertically from the surface to 100 millibars (approximately 15 km) using a  sigma-
pressure coordinate system. Air quality conditions at the outer boundary of the 36 km domain
were taken from a global model and did not change over the simulations. In turn, the 36 km grid
was only used to establish the incoming air quality concentrations along the boundaries of the  12
km grids. Table 2-7 provides some basic geographic information regarding the CMAQ domains.
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Table 2-7.  Geographic elements of domains used in RFS2 modeling.


Map Projection
Grid Resolution
Coordinate Center
True Latitudes
Dimensions
Vertical extent
CMAQ Modeling Configuration
National Grid
Western U.S. Fine Grid
Eastern U.S. Fine Grid
Lambert Conformal Projection
36km
12km
12km
97degW, 40degN
33 deg N and 45 deg N
148x112x14
213x192x14
279 x 240 x 14
14 Layers: Surface to 100 millibar level (see Table II-3)
                                      2-26

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Figure 2-10. Map of the CMAQ modeling domain. The black outer box denotes the 36 km
national modeling domain; the red inner box is the 12 km western U.S. fine grid; and the blue
inner box is the 12 km eastern U.S. fine grid.
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2.2.3  Overview of air quality using modeled and observed data
       To provide information for use in characterizing the adequacy of the current standards,
we assess the best available data for estimating the ambient concentrations of atmospheric
nitrogen and sulfur across the U.S.  Acidification and nutrient enrichment processes are largely
dependent on the cycling of total nitrogen and sulfur species.  From an atmospheric perspective,
it is convenient and consistent with current measurement and modeling frameworks to consider
the reduced and oxidized forms of atmospheric nitrogen.  Virtually all atmospheric sulfur is
considered oxidized sulfur in the forms of particulate bound sulfate and gaseous sulfur dioxide.
In order to  assess current concentrations of reactive nitrogen and sulfur, we evaluated data
available from the existing monitoring networks as well as from the CMAQ model. Regarding
the monitoring data, there are a number of important issues in understanding the measurements
of NOy provided by different monitoring networks. In principle, measured NOy is based on
catalytic conversion of all oxidized species to NO followed by chemiluminescence NO detection.
We recognize the caveats associated with  instrument conversion efficiency and possible inlet
losses.  The CMAQ treats the dominant NOy species as explicit species while the minor
contributing non-PAN organic  nitrogen compounds are aggregated.  Atmospheric nitrogen and
sulfur largely are viewed as regional air quality issues due to the importance of chemical
conversion of primary emissions into  secondarily formed species, a combination of ubiquitous
sources, particularly mobile source emissions of NOx, and elevated emissions of NOx and SO2
that aid pollutant mass dispersal and broader physical transport over large distances.  In effect,
the regional nature is due to both transport processes as well as the relatively ubiquitous nature of
sources combined with chemical processes that tend to form more stable species with extended
atmospheric lifetimes. This regionalized effect, particularly throughout the eastern United
States, dominates the overall patterns  discussed below of secondarily formed species such as
sulfate or NOy, which is an aggregate of species with the more aged air masses consisting largely
of chemically processed air dominated by secondarily formed peroxyacetyl nitrate (PAN),
particulate  nitrate and nitric acid.
       Nationwide maps of CMAQ-predicted 2005 annual  average NOy,  NHX (NH3 and NH4),
NH3, NH4,  SOx, SO4, and SO2  are provided in Figures 2-11 through 2-17 respectively. Given the
considerable gaps in air quality observation networks as discussed in the REA and ISA (EPA,
2008), modeled concentration patterns are used here to illustrate national representations of
                                          2-28

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current air quality conditions for nitrogen and sulfur. The 2005 model year reflects the most
recent available simulation for inclusion in this policy assessment. In addition, Figures 2-18 and
2-19 provide maps of 2005 annual average SO2 and SO4, respectively based on CASTNET
observations. Site specific annual average 2005 NOy measured concentrations at SLAMS
(Figure 2-20) are typically are less than 40 ppb. The spatial patterns for the 2005 modeled and
observed NOy, NHx, and 862 concentrations are similar to the 2002 CMAQ-based maps
provided in the REA, largely capturing the influence of major emissions patterns (Figures 2-2 -
2-4) throughout the nation.  The NOy patterns (Figure 2-11) reflect the distribution of NOx
emissions power generation and widely dispersed transportation sources with a spreading into
more rural locations associated with transformation of NOx to more aged NOy species such as
PAN and nitric acid, discussed in more detail in section 2-3.  Ammonia and ammonium
concentration patterns (Figures 2-13-2-14) are influenced strongly by the ammonia emissions
distribution, with marginal spreading associated with the formation of NH4.   The NHX fields are
more strongly influenced by source location, relative to sulfur, based on the fast removal of
atmospheric ammonia through deposition.  However, recent incorporation of ammonia bi-
directional flux treatment (Appendix F) does reduce NHs spatial gradients. A spreading of the
oxidized sulfur fields (Figures 2-15 to 2-17), relative to SO2, is consistent  with sulfate
transformation and associated air mass aging and transport. Note that SO2 is the dominant
contributing species in the mix of SO2 and particulate SO4with the most elevated levels in
proximity to the Ohio River valley.
                                          2-29

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            0.736637- 1.006000
            1 000001 - 3 000000
            3 000001 - 5.000000
            5 000001 - 7.000000
           | 7.000001 - 10000000
           | 10.000001 - 25 000000
           [25000001-118001373
Figure 2-11. 2005 CMAQ modeled annual average NOy (ppb; see Table 2-1 for unit
conversions).
     Legend
        0121428-1.000000
        1000001-3000000
        3000001-5.000000
        5.000001-7.000000
    ^H 7.000001 -10.000000
    ^B 10.000001 -26.524269
Figure 2-12. 2005 CMAQ modeled annual average total reduced nitrogen (NHx) (as ug/m
nitrogen - see Table 2-1 for unit conversions).
                                                  2-30

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     Legend
    NH3
        0.011665- 1 000000
        1 000001 -3000000
        3.000001 -5.000000
        5000001 -7.000000
    |^B 7000001 -10.000000
    ^B 10,000001 -24139696
Figure 2-13. 2005 CMAQ modeled annual average total reduced nitrogen (NH3) (as ug/m
nitrogen - see Table 2-1 for unit conversions).
Figure 2-14. 2005 CMAQ modeled annual average ammonium, NH4, (as ug/m N; see Table 2-1
for unit conversions).
                                             2-31

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     Legend
    SOX
       0.189676-1.665000
      | 1.665001 -3.141000
      |3.141001 -4.618000
      | 4.618001 -6.049000
      | 6.049001 -7571000
       7.571001 - 9.047000
      | 9.047001 -10.524000
      | 10.524001 -56.065598
Figure 2-15. 2005 CMAQ modeled annual average SOx, (as ug/m  S from 862 and 864; see
Table 2-1  for unit conversions).
     Legend
    SO 2
       0.131346-1.528000
       1.528001 -3.024000
       3.024001 - 4.520000
       4.520001 -6.016000
       6.016001 -7.512000
       7512001 -9008000
       9.008001 - 10.504000
       10.504001 - 12.000000
Figure 2-16. 2005 CMAQ modeled annual average SO2 (as ug/m  S; see Table 2-1 for unit
conversions).
                                                    2-32

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Figure 2-17.  2005 CMAQ modeled annual average SO4 (as ug/m  S; see Table 2-1 for unit
conversions).
Figure 2-18.  2005 annual average sulfur dioxide concentrations (total mass) based on
CASTNET generated by the Visibility Information Exchange Web System (VIEWS) (see Table
2-1 for unit conversions).
                                         2-33

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                                                                            12.9
                                                                            2.3
                                                                            1.7
                                                                            1.2
                                                                            0.5735
                                                                            O.QOOO
                                                                           uq/m3
Figure 2-19. 2005 annual average sulfate concentrations (total mass) based on CASTNET
generated by the Visibility Information Exchange Web System (VIEWS), [interpolating relative
sparse data can produce unrealistic concentration plumes as demonstrated  in the central U.S.]
(see Table 2-1 for unit conversions).
                   Annual Average NOY Concentrations (2005)
Figure 2-20. Annual average 2005 NOy concentrations from reporting stations in the Air Quality
System (AQS). (see Table 2-1 for unit conversions).
                                           2-34

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2.3    CHARACTERIZING DEPOSITION THROUGH MONITORING AND MODELS
       The removal of sulfur and nitrogen from the ambient air environments occurs through
wet and dry deposition processes. Wet deposition results from the transfer of gaseous and
particulate species into cloud droplets and their subsequent deposition as well direct scavenging
by rain and snow.  There also is deposition associated with direct contact between clouds, fog
and surfaces, referred to as occult deposition. Occult deposition is not treated explicitly in air
quality models like CMAQ and generally is assumed to be negligible with respect to
contributions to annual average total deposition, although occult deposition can provide
relatively greater contribution over shorter, episodic time frames. Dry deposition is the removal
of gases and particles from the air to surfaces, vegetation and water. The collection of rainwater
followed by chemical analysis enables direct  observation of wet deposition. Dry deposition is
not a directly measured variable in routine monitoring efforts.  It is important to pursue the
development of direct dry deposition measurements to improve model parameterizations of
deposition processes and possibly evolve into routine operations. Estimates of dry deposition
based on observations are provided through the CASTNET program.  However, dry deposition is
a calculated value represented as the product  of ambient concentration (either observed or
estimated through air quality modeling) and deposition velocity, Depf^ - vtDry • C^mb
       Deposition velocity is modeled as a mass transfer process through resistance layers
associated with the canopy, uptake by vegetation, water and soil which collectively are
influenced by micrometeorology, land surface and vegetation types and species specific
solubility and reactivity. Dry deposition is calculated through deposition velocity models
capturing these features  and using species specific ambient air concentrations.  This approach
conceptually is similar using either observed  or modeled air concentrations. Dry deposition
estimates from the Community Multi-scale Air  Quality (CMAQ) model  have been used in this
assessment to provide spatially more resolved and extensive estimates of dry deposition for
sulfur and all reactive nitrogen (oxidized and reduced) species (CASTNET does not capture
important gases such as  nitrogen dioxide, ammonia and peroxyacetyl nitrate (PAN).  All of the
relevant meteorological, land use, vegetation  and  elevation data required to estimate deposition
velocities are generated  or accessible in the CMAQ and/or meteorological pre-processors.
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       CMAQ provides a platform that allows for a consistent mass accounting approach across
ambient concentrations and dry and wet deposition values. Recognizing the limitations of
ambient air networks, CMAQ was used to estimate dry deposition to complement NADP wet
deposition for MAGIC  modeling and for the first-order acidity balance (FAB) critical load
modeling. CMAQ promotes analytical consistency and efficiency across analyses of multiple
pollutants.  EPA's Office of Research and Development  continues to enhance the underlying
deposition science in CMAQ. For the purposes of this policy assessment, CMAQ provides a
consistent platform incorporating the atmospheric and deposition species of interest over the
entire United States.  The caveats and limitations of the use of model predictions are largely
associated with the general reliance on calculated values, rather than on measurements.  Model
evaluation addressing the comparison of predictions with observed values is addressed in the
REA and summarized in Appendix F as well as a summary of ongoing and  planned model
improvements.
       CMAQ provides both concentrations and depositions for a large suite of pollutant species
on an hourly basis for 12 km grids across the continental U.S. Deposition velocities are treated
by:
       1) vdry values of gaseous pollutants are calculated in the CMAQ weather module called
the Meteorology-Chemistry Interface Processor (MCIP) through a complex function of
meteorological parameters (e.g. temperature, relative humidity) and properties of the geographic
surface (e.g. leaf area index, surface  wetness)
       2) vdry values for particulate  pollutants are calculated in the aerosol  module of CMAQ,
which, in addition to the parameters  needed for the gaseous calculations, also accounts for
properties of the aerosol size distribution
       3) vwet values are not explicitly calculated.  Wet deposition is derived from the cloud
processing module of CMAQ, which performs simulations of mass transfer into cloud droplets
and aqueous chemistry  to incorporate pollutants into rainwater.
       Due to lack of direct measurements, no performance evaluations of CMAQ's dry
deposition calculations  can be found; however, the current state of MCIP is the product of
research that has been based on peer-reviewed literature from the past two decades (EPA, 1999)
and is considered to be  EPA's best estimate of dry deposition velocities.  Although the model is
continually undergoing improvement, CMAQ is EPA's state-of-the-science computational
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framework for calculating deposition. The CMAQ was used in this assessment because it is the
state of science model for simulating sources, formation, and fate of nitrogen and sulfur species.
In addition to undergoing periodic independent scientific peer review, CMAQ bridges the
scientific and regulatory communities as it is used extensively by EPA for regulatory air quality
assessments and rules.  CMAQ provides hourly estimates of the important precursor,
intermediate and secondarily formed species associated with atmospheric chemistry and
deposition processes influencing ozone, particulate matter concentrations and sulfur and nitrogen
deposition. Simulations based on horizontal spatial scale resolutions of 12 km and 36 km were
used in this policy assessment for 2002 - 2005.
2.3.1   Current patterns of dry and wet deposition
     The National Atmospheric Deposition Program (NADP) includes approximately 250 sites
(Figure 2-21) across the U.S. providing annual total wet deposition based on weekly averaged
measures of wet deposition of nitrate, ammonium, sulfate and other ions based on the
concentrations of these ions in precipitation samples.  Meteorological models have difficulty in
capturing the correct spatial and temporal features of precipitation events, raising the importance
of the NADP as a principal source of precipitation chemistry. The NADP has enabled several
organizations to participate in a measurement program with a centralized laboratory affording
measurement and analysis protocol consistency nationwide. Virtually every CASTNET site is
located at an NADP site and the combined NADP/CASTNET infrastructure is a starting point for
discussions addressing future NOx and SOx monitoring needs.  Analysis of organic bound
nitrogen recently has been added to the NADP suite of parameters. Consideration might be
given to adding NADP sites in locations where ambient air monitoring is conducted to assess
compliance with a secondary NOx/ SOx standard. For consistency, we use  CMAQ developed
national maps of total  deposition.  Additional NADP maps  of wet deposition are available at
http://nadp.sws.uiuc.edu/.
                                          2-37

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                                    Ion      deposition,
                                                                           Ammonium as NH4*

                                                                                 sOJ
      Sites not pictured:                                 '  ,   ".*   --!,''   '              1.0-1.5
      VI01     0.4 kg/ha           "I            '•   .         • "          ,  ,VT         2.0-2.5
                u                       ?,,                           ='  '          2.5-3.0
                                        •..-""                       •     -         3.0-3.6
                                        ' •'                                        3.1 • 4.0
                                                                                 4.0 - 4.S
                                                                                 > 4 S
  National Atmospheric Deposition Program/National Trends Network
  hWp/,'nadp.sws uiucedu
Figure 2-21.  Location of approximately 250 National Atmospheric Deposition Monitoring
(NADP) National Trends Network (NTN) sites illustrating annual ammonium deposition for
2005.  Weekly values of precipitation based nitrate, sulfate and ammonium are provided by
NADP.
2.3.2  Characterizing deposition through CMAQ
       Total deposition for nitrogen, reduced nitrogen, the ratio of reduced to total nitrogen and
sulfur (Figures 2-22 and 2-25) basically follow the patterns of ambient air concentrations
described earlier. The contribution of reduced nitrogen to total nitrogen deposition (Figure 2-24)
illustrates the strong influence of agricultural based ammonia emissions, particularly in upper
midwest and eastern North Carolina. These maps represent the deposition values used in the
calculations of areas likely not meeting alternative standards as decribed in in section 7.5.
                                           2-38

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     Legend
        0.675269-2.000000
        2.000001 -3.000000
        3000001 -4.000000
        4.000001 -5.000000
        5.000001 -7.000000
        7.000001 -9.000000
        9.000001 -14 000000
        14.000001 -19.189981
Figure 2-22. 2005 CMAQ modeled oxidized nitrogen deposition (kgN/ha-yr).  (see Table 2-1
for unit conversions).
     Legend
    • 0.113673-2.000000
    • 2.000001 -3.000000
       3.000001 -4.000000
       4000001 -5.000000
       5000001 -7000000
       7000001 -9.000000
    • 9.000001 -14.000000
    • 14.000001 -20.000000
Figure 2-23. 2005 CMAQ modeled reduced nitrogen deposition (kgN/ha-yr).  (see Table 2-1  for
unit conversions).
                                                     2-39

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   Legend
  «DN_to_TOTN_ratio
     0086774-0.180000
   | 0.180001-0.292000
   • 0282001-0.384000
   | 0384001-0.486000
   • 0486001 -0588000
     0588001 -0.690000
   • 0.690001 - 0.793000
   B 0.793001 -0886040
Figure 2-24.  2005 CMAQ modeled ratio of reduced to total nitrogen deposition.
   Legend
  rD_S_TOT
  ^B 0.240733 •
  ^f 1.000001 •
     2.000001 •
     3000001 •
     6.000001-
     10.000001
  HI 16 000001
  ^•24 000001
     30.000001
1.000000
2.000000
3.000000
6000000
10.000000
• 16.000000
• 24.000000
• 30.000000
-89142693
Figure 2-25.  2005 CMAQ modeled oxidized sulfur deposition (kgS/ha-yr).  (see Table 2-1  for
unit conversions).
                                                        2-40

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2.3.3  Relationships between patterns of ambient concentrations and deposition
       The development of an aquatic acidification standard relies on relationships between air
concentrations and deposition. Consequently, it is informative to understand some of the basic
patterns and relationships between concentrations and deposition of SOx and NOy species.
While there are the obvious first order associations that we see between concentration patterns
(Figures 2-1 1 to 2-17) and deposition (Figures 2-22 to  2-25), as well between emissions and
concentrations, there exist marked differences between concentration and deposition at the
individual species level. While the differences between emissions and air concentrations can
generally be attributed to a plethora of atmospheric chemistry and transport mechanisms that
change the nature and location of emitted species, the differences between concentration and
deposition are all about the inherent characteristics of each species and how various
meteorological and surface attributes (meaning landtypes, water systems, vegetation, suspended
cloud and rain droplets) influence the transfer of a species to a (or through and within) a surface.
This section describes these relationships and provides background for the discussion on the
selection of ambient air indictors (section 7.1).
     species
       Air quality models and deposition models that use direct observations calculate
deposition on a species by species basis to account for differences in deposition velocities.
Consequently, the relative fractional contributions of individual NOyor SOx species to
deposition or concentration is influenced by the differences in species deposition velocities. For
example, nitric acid with a high deposition velocity would exhibit a larger relative contribution
to overall deposition compared to ambient concentrations in a particular area (Figures 2-26 and
2-27).  The dominant ambient air NOY species are NO, NO2, F£NO3, P-NO3 and PAN.  Near
source urban environments typically have a  relatively higher fraction of NOx (NO and NO2)
compared to the products of NOx reactions,  nitrates and PAN, which are relatively more
dominant in rural locations (Figures 2-27 - 2-30).
Sulfur Species
       The use of SO2 and SO4 does reflect the use of individual where it is practical to measure
each species independently. Although sulfur dioxide and particulate sulfate contribute
approximately 60 and 40 %, respectively, to ambient SOx concentrations, sulfur dioxide is the
                                           2-41

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dominant contributor to SOx deposition (Figure 2-31), which is consistent with CASTNET
observational studies (Sickles and Shadwick, 2007). With minor exceptions, the PM2.5 fraction
generally accounts for over 80% of the ambient sulfate mass. However, as particle size diameters
increase beyond 2.5 |i,  gravitational settling imparts greater influence resulting in substantially
enhanced deposition velocities.  Consequently, the sulfate mass in size fractions greater than 2.5
|i potentially provides correspondingly greater contribution (to as much as 50% of dry sulfate
deposition in certain locations (EPA , 2008; Grantz et al, 2003), which has implications for
monitoring that are discussed below.
2.3.4  Monitoring Considerations
       The differences in the relative patterns between ambient air and deposition on a species-
by-species basis illustrate a number of challenges and considerations in developing a monitoring
strategy.  It is clear in the Adirondacks and Shenandoah areas, for example, that nitric acid is the
most dominant contributing species  from a deposition perspective (Figure 2-26), with significant
contributions from particulate nitrate, PAN and NC>2.  The original source of emissions (NO
accounts for 90-95% of all emitted NOx) ultimately is transformed in the atmosphere and
provides very small faction of oxidized nitrogen in ambient air and deposition in rural
environments.   The combination of nitric acid and particulate nitrate consistently contribute
greater than 50% of the oxidized nitrogen dry deposition load, whereas PAN and NC>2 contribute
roughly 15-25% of the deposition load.
       Dry deposition of NOy is treated as the sum of the deposition of each individual species
in advanced process based air quality models like CMAQ.  This raises the question of the
relative importance of acquiring individual species measurements to a single aggregated
measure, NOy. For example, individual measurements of the dominant NOy species (HNOs,
particulate nitrate, NO2, NO, and PAN) could be coupled to their distinct deposition velocities to
estimate  dry deposition and provide useful diagnostic information to improve characterization of
deposition processes.  Currently, technology for measuring NO2 in rural locations, HNOs, and
PAN generally is not available for routine network applications. If certain species provide
negligible contributions to total NOy deposition, then perhaps they could be excluded for the
purpose of deposition assessments.  All of the nitrogen species that constitute NOy have species
specific dry deposition velocities.  Species with especially low relative deposition velocities,
such as nitrogen dioxide, may contribute insignificant amounts of deposition relative to species
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with high deposition velocities such as nitric acid.   Based on the reasoning that a larger fraction
of the deposited NOy is accounted for by total nitrate combined with the availability of reliable
total nitrate measurements (the sum of nitric acid and particulate nitrate) through CASTNET, a
total nitrate measurement may be adequate for deposition based assessments.
       These patterns suggest the possibility of using total nitrate as a key indicator for
acidifying deposition contributions associated with oxides of nitrogen.  However, a nitrate
observation alone would miss a considerable fraction of the ambient NOy burden reflected in
significant levels of NO2 and PAN.  Characterization of NC>2 deposition is an area requiring
further refinement especially considering that NC>2 is a significant component of total oxidized
nitrogen. Zhang et al. (2005) suggest that NC>2 contributes up to 36% of dry NOy  deposition in
rural Eastern Canadian locations, and suggest, based on observational evidence (Figure 2-27),
that in some locations NC>2 deposition may be similar to nitric acid contributions.
       Another way of addressing the relative benefit of using part of the NOy mix relative to
total NOy in regard to deposition is to probe the dynamic response of changes in oxidized
nitrogen deposition to changes in ambient concentrations NOy and nitric acid.  Dynamic
response refers to sensitivity of the ambient to deposition response with respect to changes in
NOx emissions, which is relevant to air quality management as ambient indicators are used to
assess if an area meets  or exceeds a target value in current and future time frames. While such a
response to emission changes may be linear or non-linear, the details of which are encoded in
chemical transport air quality models like CMAQ, typically there is a directional relationship
between the change in the precursor emissions and the target species of interest.  By extension,
one would expect that a significant change in emissions of NOx would lead to a change both in
the ambient and deposition fields of NOy species,  recognizing that NOy species all evolve from
NOx emissions, which is dominated by nitrogen oxide, NO.  We can apply this reasoning to the
consideration of using HNOs as a more narrowly defined indicator, relative to NOy. A 2005
base case and projected 2030 CMAQ simulation, with roughly 50% NOx and SOx reductions,
respectively, are used to illustrate the relationship of HNO3 and NOy concentration changes to
changes in NOy deposition (wet and dry) which address the question:  Does the indicator
respond in a manner directionally similar to deposition over periods of significant emissions
reductions?  Based on this paired set of current and future projection scenarios, the absolute
change in ambient NOy is greater than the absolute change in nitric acid concentrations and both
                                           2-43

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NOy and nitric acid respond in a similar relative directional manner (Figure 2-32).  The higher
magnitude of absolute change is a desired attribute from an indicators perspective as signal
changes over time are more likely to be detectable.
       These examples suggest that the acidifying contributions of all NOy species should be
accounted for in linking ambient air to deposition.  Ideally, observations of individual NOy
species are preferable as they allow for a more refined understanding of the contribution of
individual species to deposition, and afford data to diagnose air quality model behavior that can
lead to improved parameterization of deposition processes. However, limitations of available
technology suggest that measurements for aggregated NOy are available for routine application.
An aggregate NOy measure does,  in concept, capture the potential for acidifying contributions of
all oxidized nitrogen species. Nevertheless, complementary measurements of NO2, HNOs, p-
NOs and PAN to allow for diagnostic evaluations of both air quality models and the NOy
measurement itself should be strategically placed in two to five areas, in different air quality
mixes and ecologically relevant locations.
       Measurement technology issues generally are not as complex for SOx as they are for
NOy and individual NOy species,  partly because just two sulfur species, sulfur dioxide and
particulate sulfate, dominate oxidized sulfur composition in the atmosphere.  However, as noted
earlier there are concerns related to capturing the full range of sulfate particle size fractions.
       Ammonia and ammonium  ion both provide the potential to contribute acidifying
deposition and, therefore, should be accounted for in assessments addressing acid deposition.
Characterization of reduced nitrogen deposition processes is an active developmental area which
would benefit markedly from NHx measurements in order to assess modeled predictions of
ambient patterns of ammonia and ammonium. This need for monitoring ammonia in rural
environments is further supported by emerging evidence that ammonia acts as a regionally
dispersed species  based on the inclusion of ammonia bi-directional flux in CMAQ simulations as
discussed  in Appendix F and Dennis et al, 2010.  Monitoring method approaches under
consideration for routine application typically are limited to time averaged filter and denuder
technologies, including passive sampling approaches which are utilized in the new NADP
AMON network.
       As discussed earlier,  two - five locations nationally, in airsheds with different
atmospheric chemistries, that sample not only for the NAAQS indicator NOy but for the suite of
                                          2-44

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major NOy species as well; HNOs, p-NOs, PAN, NO2, and NO as discussed earlier. Not only is
this important from a modeling and process diagnosis perspective, but it is especially useful in
the introduction of new measurements that have a limited track record to provide insight into
instrument performance. In the case of NOy, it is even more relevant since there effectively are
no standards that explicitly challenge instrument accuracy given the highly variable nature of
NOy species distribution and the instability associated with mixing NOy gases.  This quality
assurance issue is analogous to PM2.5 where aerosol standards are not available and measurement
accuracy is judged against periodic challenges relative to a "gold standard" instrument. Reduced
nitrogen measurements of ammonia and ammonium ion are recommended at all locations with
FRM/FEM instruments based on the need to support the AAPI as discussed above.
Sampling frequencies
       The averaging time for the standard is likely to be an annual average,  perhaps based on 3-
5 years of data collection to minimize the influence of interannual variability in meteorology,
especially precipitation. Conceptually, extended sampling periods no longer than one year would
be adequate for the specific purposes of comparison to a standard.  However, there are
significant peripheral benefits relevant to improving the scientific foundation for subsequent
reviews and a variety of related air quality and  deposition assessments to be gleaned from more
highly time resolved data. In particular, the critical role of air quality models in deposition
assessments implies value to be derived from measurements that support model evaluation and
improvement. Many of the monitoring approaches that are used throughout the nation sample
(or at least report out) on daily (PM2.5 chemical speciation), weekly (CASTNET) and hourly (all
inorganic gases) periods. There is a tradeoff to consider in sampling period design. For
example, the weekly CASTNET collection scheme covers all time  periods throughout a year, but
only provides weekly resolution that misses key temporal and episodic features valuable for
diagnosing model behavior. The every third day, 24-hour sampling scheme used in IMPROVE
and EPA speciation monitoring does provide more information for a specific day of interest yet
misses 2/3 of all sampling periods. The missing sampling period generally is  not a concern when
aggregating upward to a longer term average value as the sample number adequately represents
an aggregated mean value. Additionally, there is a benefit to leveraging existing networks which
should be considered in sampling frequency recommendations.  A possible starting point would
be to assume  gaseous oxidized species, NOy and SO2, are run continually all year reporting
                                          2-45

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values every hour, consistent with current routine network operations. Sulfate sampling periods
should coincide with either the chemical speciation network schedules or with CASTNET.
There are advantages to coordinating with either network. Ammonia gas and ammonium ion
present challenges in that they are not routinely sampled and analyzed for, and the combined
quantity, NHx is of interest. Because NHx is of interest, some of the problems of volatile
ammonia loss from filters may be mitigated. However, for model diagnostic purposes,
delineation of both species at the highest temporal resolution is preferred.
       Sample collection period is not an issue for gaseous measurements of NOy and SO2 that
operate continuously.  However, consideration should be given to using the CASTNET filter
pack (FP) for SC>2 measurements to maximize leveraging of monitoring assets, assuming the FPs
will be used for particulate sulfate.  However, the availability of highly time resolved data will
support the continual evaluation of 862 and sulfate balance in air quality modeling systems
which is  a critical underpinning  for both human and ecosystem health assessments.
Spatial scales
      The current observation network for NOy, NHX and SOx is very modest and includes a
monitoring network infrastructure that is largely population oriented with the exception of
CASTNET and IMPROVE. While there is platform and access infrastructure support provided
by CASTNET, NADP and IMPROVE, those locations by themselves are not likely to provide
the needed spatial coverage to address acid sensitive watersheds across the United States.
Ambient monitoring at every watershed will not be required given the reality of resource
constraints and the relative spatial homogeneity of air concentrations that are averaged over
annual time periods  and within 'acid sensitive" areas. The spatial monitoring requirements will
be associated with the  determination of acid sensitive areas, which is discussed in chapter 7.
The number of sites  per area will be addressed in rule development and general guidance based
on an understanding of the spatial variability of NOy, NHX, sulfate and SO2 combined with
resource  allocations will help inform those decisions.
      Critical load  models applied for the purposes of this standard would be based on annual
averages, which would effectively serve to dampen much  of the spatial variability.  Furthermore,
the development of an area-wide depositional load tradeoff curve implies focus on region wide
characterization. Toward that end, CMAQ concentration fields will provide insight into the
likely spatial representativeness  of monitors leading to efficient application of monitoring
                                          2-46

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resources. For example, the CMAQ based spatial coefficient of variation (standard
deviation/mean) of oxidized nitrogen in the Adirondacks was 1.46%. Improved dry deposition
estimates will result from enhancements of ambient monitoring addressing the N/S secondary
standards as each additional location could serves a similar role that existing CASTNET sites
provide in estimating dry deposition.
Candidate monitoring methods
       Ambient NOy, 862 and paniculate sulfate (SO/t) concentrations are likely candidates for
ambient air indicators (section 7.1). All of these indicators are measured in different places
within the current routine monitoring networks.   Traditionally, Federal Reference or
Equivalency Method (FRM/FEM) status of measurement techniques is used for estimating air
concentrations for NAAQS comparisons. A FRM for SC>2 exists, but not for NOy or SO/t.  Only
recently have NOy measurements, which historically were viewed as research venue
measurements, been incorporated as "routine" observations,  partly as a result of the NCore
program.  Paniculate SO4 is measured at over 500 sites nationally, and there is a general
consensus that methods available are reliable and provide consistent data.
       Particulate-SO4. Particulate sulfate (p-SO4) has been measured for several years in the
IMPROVE,  CASTNET and EPA CSN networks. The nation has over 500 24-hour average,
every third day sulfate measurements produced by the PM2.5 speciation networks (IMPROVE
and EPA CSN) and nearly 80 CASTNET sites that provide continuous weekly average samples
of sulfate with an open inlet accommodating all particle sizes.  As discussed above, particle size
diameters increase beyond 2.5 ji should be accounted for in deposition based assessments,
perhaps ruling out the use of PM2.5 data serving as indicators for a NOX/SOx secondary
standard.
       The routinely operating methodology for p-SO4 is based on an  integrated (i.e., time
averaged over several hours or days) sample collection on a Teflon filter followed by ion
chromatography (1C) detection in the laboratory.  Two major variations of this approach are
applied in the PM2.5 speciation (exclusion of particles larger than 2.5 |i and 24-hour collection
typically every third day) and CASTNET (weekly average integrated sampling all year with an
open inlet to include all size fractions). There are additional variations related to inlet design and
flow characteristics of PM2.5 speciation samplers in which two designs are prevalent in the
networks: (IMPROVE and EPA CSN  SASS samplers).  These variations are considered minor as
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sulfate species (dominated by ammonium sulfate) typically are not subject to major sampling
artifacts associated with volatilization or condensation.   The difference in inlets (open vs. 2.5 ji)
is perceived by some as not an issue of concern as 80 - 90 % of the PM sulfate mass is distributed
in size fractions less than 2.5 ji.  However, the higher deposition velocities associated with larger
diameter particles argue for including all size fractions as discussed above.   Continuously
operating in-situ sulfate instruments that allow for hourly, or less, data reporting are available.
However, the limited deployment (less than 20 sites nationally) of these instruments combined
with the 2.5 |i inlet cutoff configuration preclude consideration at this time.
       The CASTNET FP offers three important attributes: a history of high quality data,
existing infrastructure and network to build on and an open inlet to capture the full range of
particle diameters. EPA intends to develop FRM status for this method.  A significant
additional advantage of using the FP method will be the availability of important co-measured
species (e.g.,  862, total nitrate, and ammonium).  While EPA plans to expedite the certification
process for the CASTNET FP, in the future consideration should be given to other available
methods to more efficiently leverage network assets. For example, the SASS sampler potentially
would accommodate ammonia gas and ammonium ion measurements, as well as other standard
chemical speciation parameters depending on the configuration of this multi channel system.
Continuous sulfate measurements would be extremely useful for model evaluation, especially
considering the availability of continuous 862 data that would be required as part of the NAAQS
indicators.  A performance based approach to meet equivalency requirements, given the variety
of sulfate measurement approaches and well vetted and accurate analytical procedures.
       SO2. A FRM is available for SO2. See 75 FR at 35554-56 and 35593-95 (June 22, 2010)
(adopting a second FRM for 802).  As part of the NCore network development effort, trace gas
SC>2 analyzers capable of sub ppb resolution became commercially available and are the
preferred instruments for implementation in rural locations.  As discussed above, the near
continuous data output of gaseous  analyzers is desired for peripheral support of model
evaluation.  Nevertheless, the convenience and resource savings associated with the CASTNET
FP suggest that Federal Equivalency Method (FEM) status should be incorporated in concert
with the sulfate certification process.
       NOy.  In principle, measured NOy is based on catalytic conversion of all oxidized species
to NO followed by chemiluminescence NO detection. While there are caveats associated with
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instrument conversion efficiency and possible inlet losses, the technique is considered adequate
and routinely operational. Approximately 25 sites (out of a planned 75) in EPA's NCORE
network are operating NOy instruments, and additional sites are operated in SEARCH,
CASTNET and other programs. NOy measurements are nearly continuous, reporting at hourly
intervals providing far greater temporal information compared to filter or denuder based
methods.
       FRM certification for NOy presents challenges given the limited history of routinely
operating instruments. NOy measurements are in a transition period from largely being viewed
as a research level measurement to now being deployed as a routine measurement in EPA's
national 75 site NCORE network.  The general consensus on NOy measurement is that the
methodology is sound and applicable for routine/regulatory use, but there does not exist a well
defined understanding of the quality of NOy data. Inorganic dry nitrate (nitric acid and
particulate nitrate) is measured routinely in the CASTNET network with filter packs (FP).
Acquiring FRM status for NOy instruments may require better characterization of the conversion
efficiencies, mass loss and updated guidance on operating and siting procedures.
       One of the challenges associated with specifying performance attributes for p-SO4 and
NOy is the lack of specific challenge standards. For example, instruments measuring discrete
gases such as ozone or nitrogen oxide can be challenged by comparing an instrument's reading
when measuring known concentrations of gases which are readily provided for single gas
concentrations. Particle standards are not available. NOy performance typically is challenged by
known mixtures of NO2, and  occasionally with N-propyl nitrate, which only addresses part of the
spectrum of nitrogen species in an NOy mix. Consequently, instrument performance in EPA's
national networks for aerosol mass is quantified in terms of bias and precision relative to a co-
located "performance evaluation" instrument.  There is no comparable program in place for p-
SO4orNOY.
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                     Mean Species Breakdowns aTN Concentrations
                     Mean Species Breakdowns of N Depositions
                                                                 PAN
                                                                 PANX
                                                                 NTR
                                                                 NO
                                                                 NO3
                                                                 NO2
                                                                 N205
                                                                 MONO
                                                                 HNO3
                                                               DDEPJJ03
                                                               DDEP.PANT
                                                               DDEP.NTR
                                                               DDEP_HONO
                                                               DDEPJHN03
                                                               DDEP_N205
                                                               DDEP_N02
                                                               DDEP_NO
Figure 2-26. Annual 2002 - 2004 CMAQ derived annual average fraction of ambient
concentrations (above) and dry deposition (below) of individual NOy species delineated by the
Adirondack and Shenandoah case study areas and the remainder of the Eastern U.S. domain.
NTR refers to non-PAN organic nitrates.  PANX refers to aggregation of PAN  type compounds,
other than PAN, specifically.
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         Jan  Fed Mar  Apt  May Jun  jul  Aug  Sep OcJ  Nov Dec

               Note Oct and Nov NO. valu«»r9«tinia1ed
                   M»  Mly Jun  Jill  AtlQ  Sep Ocl N»v OfC

                 Jan and R* PAN valu«s an osbinatKj
Figure 2-27.  Examples of the Relative Abundance of Several NOy Species Measured at Two
Rural Southeastern Canadian Sites as a Fraction of the Total Measured NOy Concentration —
Kejimkujik, NS, (top) and Egbert, ON, (bottom) during 2003. Although both sites are in rural
locations, the Kejimkujik, NS site represents more aged air masses as it lies considerably further
downwind from major sources of NOx relative to the Egbert  site. (Source: NARSTO, 2011).
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      Legend
      N02 to NOy Ratio
          0.136-0.184
      ^H 0.185 -0.306
      ^H 0.307-0.401
      ^H 0.402 - 0.466
      ^B 0.467 - 0.547
          0.548 - 0.637
      ^•J 0.638 - 0.728
      ^M 0.729 -0.819
Figure 2-28.   Annual average fraction of NOy ambient air contributed by NC>2 based on 2005
CMAQ Eastern U.S. simulation at 12 km grid cell resolution.
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       Legend
       HNO3 to NOy Ratio
             0.006 -0.051
         !• 0.052 - 0.098
         ^| 0.099 -0.144
         •• 0.145-0.190
         ^HO 191 -0.237
             0.238 - 0.283
         H 0.284 - 0.330
         ^H 0.331 - 0.407
Figure 2-29.   Annual average fraction of NOy ambient air contributed by HNOs based on 2005
CMAQ Eastern U.S. simulation at 12 km grid cell resolution.
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      Legend
      PAN to NOy Ratio
            0.002 - 0.038
       ^H 0.039 - 0.073
       ^| 0.074 -0.108

       ^H 0.145-0.179
            0.180-0.214
       ^H 0.215 -0.249
       ^B 0.250 - 0.268
Figure 2-30.  Annual average fraction of NOy ambient air contributed by PAN based on 2005
CMAQ Eastern U.S. simulation at 12 km grid cell resolution.
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                      Mean Species Breakdowns of S Concarlratiars
                                                                   S04
                                                                   SO2
                      Mean Species Breakdowns o' SDflposltlone
                                                                 DDEP_S04
                                                                 DDEP_S02
Figure 2-31.  Annual 2002 - 2004 CMAQ derived annual average fraction of ambient
concentrations (above) and dry deposition (below) of individual SOx species delineated by the
Adirondack and Shenandoah case study areas and the remainder of the Eastern U.S. domain.
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S-   '
TJ
o
(jH
                                                                      it*:
                                                        •_.-•

                                                    o   o
               HN03change (ppb)
 OQ
  ro
    - -r -
                                                           SI      3]

                                                                  '•-•
          •S     -5    •*    -3    •?

               NOychange (ppb)
Figure 2-32.   Relationship of the change in total oxidized nitrogen deposition to change in
ambient nitric acid (top) and ambient NOy (bottom) based on changes in concentration and
deposition fields associated with current (2005) and reduced emission CMAQ simulations.
The values are based on the changes imparted for each 12 km grid cell within the Adirondack
region.  The NOx and SOx emissions reflect reductions of 48% and 42%, respectively, across the
entire Eastern U.S.  The left side panels reflect absolute differences (reduced - 2005) and the
right hand side reflects relative changes (2005 - reduced)/2005.
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2.4    CHARACTERIZING FRESHWATER AQUATIC SYSTEM CHEMISTRY USING
       MODELS AND MEASUREMENTS
       This section introduces basic water chemistry concepts and soil and watershed processes
incorporated in biogeochemical models used to estimate changes in water quality driven by
atmospheric deposition of nitrogen and sulfur. A summary of the of those models, with an
emphasis on those used in this assessment, as well as the major monitoring networks providing
water quality data relevant to acidification of freshwater systems are included as a reference
source to allow for a more focused policy relevant discussion of the standard in chapter 7.
2.4.1   Water chemistry basics
       Throughout this document basic water chemistry parameters and concepts such as pH,
acid neutralizing capacity (ANC), dissolved aluminum and charge balance are incorporated in
much of the rationale applied in developing the form of the standard. A brief discussion of
chemical equilibrium and electroneutrality and acid-base chemistry provide a basis for
understanding much of the terminology and rationale relevant to aquatic acidification.  Much of
the basic descriptions here are based on Stumm and Morgan (1981), which more or less has
served as the guide to aquatic chemistry of natural systems.
       First, neutrality is always adhered to, meaning that the sum of positively charged cations
equals the sum of positively charged anions.  Natural water systems are dominated by
substances which dissociate into ions (cations and anions) that are held to two conditions.   First,
the degree of dissociation is governed by the equilibrium relationship of the dissociated ions and
its "parent" compound.  In a solution of pure water, water dissociates into hydrogen ion, H+, and
hydroxide ion, OH-. The dissociation constant of water, Kw = 1*10"14,  at standard temperature
and pressure conditions, which is expressed as:
       [OH][H] = KW=1*10"14                                             (2-1)
Equation 2-1 describes the equilibrium condition.  The second condition to be adhered to is
electroneutrality.  The only species in solution is pure water, H2OH2), and it ion, H+ and OH".
Consequently, the concentration of OH" must equal the concentration of H+ to maintain a neutral
solution. Therefore, the concentration of OH is the as H = l*(10)-7 for a solution of pure water.
       The variable, pH, refers to the negative logarithmic value of hydrogen ion concentration.
This explains why solution of pure water has a pH value of 7.  So, two fundamental principles of

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water chemistry, equilibrium and charge neutrality essentially determine the chemical
characterization of natural water systems.
       Principles of acid-base chemistry, which builds on electroneutraility and equilibrium
concepts, helps explain the meaning of acidification and ANC.  Although various definitions of
acids and bases exist, the most relevant explanation builds off of the Bronsted concept that an
acid is a substance that can donate a proton to another substance and a base can receive a proton.
Hydrogen ion essentially can be thought of a proton,  as indicated by the positive charge,
although the H+ symbol really reflects a hydronium ion, H3O+, where water acts as an acid that
donates a proton in the form of H3O+ which for convenience is symbolized as H+. Acids are
substances that donate more protons than they receive relative to a reference substance, which
for convenience is water. A strong acid such as hydrochloric acid HC1 dissociates nearly
completely into hydrogen and chlorine ions with an equilibrium relationship defined by:
       Ka = [H+][Cr]/[HCl] = 10+3 at 25°C                                          (2-2)

If a known amount of HC1, HClt is added to water, electroneutrality (also referred to as the
proton condition) must be maintained, meaning that:
                    +
       OH' + Cr = H+;                                                             (2-3)
In addition, the dissociation of water as described above must be maintained:
       [OH"][H+] = 10"14;                                                           (2-4)
And Cl mass also must be retained:
       [HC1] + [Cr] = [HCl]t                                                       (2-5)
       Consequently, for a known amount of HClt added to pure water, equations 2-1 to 2-4 are
easily solved as there are four equations and four unknowns, [HC1], [OH"], [H+] and [Cl"].
Because the addition of HC1, which is a strong acid, results in the addition of strong anions, Cl",
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the solution maintains neutral charge by the addition of hydrogen ions and consequently the pH
is lowered when a strong acid is introduced.
       A strong base added to water must adhere to the same equilibrium and chare neutrality
conditions as described above.  For example, the addition of a known amount of strong base,
sodium hydroxide, NaOHt, dissociates into Na+ and OH- and adheres to the equilibrium
condition:

       Kb = [Na+][OH']/[NaOH]                                                    (2-6)

Electroneutrality is met by:

       [Na+] + [H+] = [Off],                                                        (2-7)

resulting in a reduction of H+, and raising of pH, associated with the addition of a strong cation,
Na+.  These concepts that associate strong cations with bases and strong anions with acids
explain much of the formulation of ecosystem water chemistry models that balance the
acidifying atmospheric deposition of strong anions (NXV and  SO/f2) with the natural supply of
strong cations (Ca+2, Mg+, K+, Na+). The term strong, broadly represents the near complete
dissociation of the parent salts or acids/bases from which the ions are derived from.
       Natural aquatic systems are more appropriately explained as a system of weak acids that
do have large dissociation constants.  The examples of strong acids and bases above illustrate
basic acid base and water chemistry equilibrium concepts and establish some context for
explaining titration and ANC a little later in this discussion.  The most  common acid in natural
systems is a weak acid, carbonic acid, H2CO3, which originates is formed from the reaction with
water of dissolved atmospheric CC>2.   Carbonic acid exists in equilibrium with its dissociated
ions, bicarbonate, HCCV, and carbonate, C(V2.  Therefore, to meet the electroneutrality
condition, hydrogen ion must now balance additional weak anions in addition to hydroxide ion,
       [H+] = [Off] + 2[CO3'2] + [HCCV];                                           (2-8)
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To the extent carbonic acid dissociates (Figures 2-33 and 2-34), it adds negatively charged
anions that can only be balanced by hydrogen ion, consequently lowing the pH of a system of
pure water exposed to atmospheric CC>2.  This example explains why natural water often has a
pH lower than 7 of approximately 5.7.
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  pC
                                                          11
Figure 2-33.  Equilibrium diagram illustrating distribution of carbonate species as a function of pH which is closed
to atmospheric CO2 exchange and therefore has a constant fixed amount of dissolved carbonate.  The intersection
where H4" equals HCO3" is the common equivalence point, approximately a pH of 4.5, used when titrating a solution
with strong acid to determine ANC. The amount of strong acid that it would take to reach that equivalent point is a
measure of ANC or alkalinity. Adopted from Stumm and Morgan, 1981.
        -1
        -2
    c
    .o
    '
        -4
    o

    §   i
    o>
    °   -6
        7

        -8

        -9
b             Open system,
constant CO2 partial pressure of 10~3 5 atm
                                                             10   11
                                             pH
Figure 2-34. The same system open to atmospheric CO2 exchange where the amount of dissolved carbon is
determined by the partial pressure of atmospheric CO2 and pH.  Note that a pH of about 5.7 reflects a pH of pure
water exposed just to atmospheric CO2.
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Natural water systems are never as simple as just consisting of water and carbonic acid and
carbonate ions, as typically there is a natural supply of strong base cations, CB, due to weathering
of rocks and soils, atmospheric deposition, and decomposing biomass. There also is a supply of
strong anions, CA (NO3" and S(V2), derived mostly with atmospheric deposition and nitrification
processes in vegetation and soils.  For a natural carbonate system in the presence of strong
cations and anions, electroneutrality is given by;
CB + [ H+] = [Off] + 2[C03-2] + [HC03-] + CA;                                       (2-9)
Therefore:

CB - CA =  [OH'] + 2[CO3'2] + [HCO3~] - [ H+].                                       (2-10)

       Defining ANC.  The right side of equation represents the capacity of the system to
neutralize available excess protons, H+, and conceptually represents acid neutralizing capacity,
ANC.  The actual definition of ANC is more specific and is based on how much strong acid it
takes to titrate a solution to a defined reference or equivalence point. In effect, the equivalence
point is synonymous with the point where there no longer is a deficit of protons. This definition
of ANC reflecting the difference between major cations and anions (CB - CA ) is operationally
defined by equation 2-10 and is  incorporated in many ecosystem models and this policy
assessment:
ANC = 2([CA2+] + [Mg2+]) + [K+] + [NH4+] - (2[SO4 2"] + [MV] + [Cl'])               (2-11)
Two other related terms used in water quality are acidity (ACY) and alkalinity (ALK).  Acidity
can thought of as the opposite of ANC (e.g., ACY = CA - CB), and reflects the excess of protons.
From a titration perspective, ACY is defined as how much strong base it takes so that no
protonation exists. Alkalinity is very similar to ANC and some of the differences in definitions
regarding both ANC and ALK are operationally defined as explained in Hemond, 1990. To
some extent, alkalinity reflects just the carbonate component of ANC, whereas ANC accounts
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more fully for other weak organic acids.  In this assessment, we emphasize ANC and the
occasional use ALK is intended to reflect ANC conditions and meaning.
      pH and ANC relationships.   ANC is a conserved property. This means that ANC can be
tracked in a mass balance sense as the level of ANC in a system (e.g., a lake or stream) is
calculated by adding how much ANC initially  exists with how much flows in and is deposited,
balanced how much flows out.  The term "mass balance" underlies the basic formulation of any
physical modeling construct, and refers to the accounting of the flow of mass into a system,  the
transformation to other forms, and the loss due to flow out of a system and other removal
processes.  Hydrogen ion is not a conserved property as its concentration in a system is not
related to the inflow and outflow of hydrogen ion, but influenced by several factors such as
temperature, atmospheric pressure, mixing conditions of a water body and the levels of several
chemical species in the system which all exist, or at least move towards, a state of equilibrium.
The conservative nature of ANC also can be explained by Equation 2-11  in which the quantities
of strong cations and anions are directly attributable to inputs to and outputs from a system.
Strong cations (Mg+, Ca+2, K+, Na+) and strong anions (NO3, SO4, Cl) are always completely
dissociated in surface waters, that is why they are referred to as strong ions. Consequently, they
can be accounted for in basic modeling approaches. Hydrogen ion, on the other hand,  is
dependent on the balance of all ions in meeting electroneutrality conditions, as are other "weak"
ions  associated with dissolved inorganic (DIG) and organic (DOC) carbon.
       The only condition that is always held constant for hydrogen ion is its relationship with
hydroxide ion, OH-, where the product of hydrogen ion and hydroxide ion concentrations always
= 1 xlO"14 eq/L at standard conditions, which reflects the equilibrium relationship between water
and its hydrolysis products, OH" and H+.  The addition of acids (strong anions) or bases (strong
cations) changes the concentrations of hydrogen ion. Also, changes in temperature effect
hydrogen ion concentration. Because of the influence on hydrogen ion of equilibrium constraints
and other factors constraints, the concentration can be highly variant and  not modeled as direct
function of accounting for hydrogen ion supply and removal. That does not mean that hydrogen
ion cannot be modeled, as water chemistry models calculate pH by solving for the total charge
balance in the system while accounting for the equilibrium relationships of weak acids and
adjusting for temperature. The amount of dissolved inorganic carbon and ANC basically
determine pH. Because of these dependencies, the response of hydrogen ion to acidification
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inputs is inherently nonlinear.  pH measurements themselves are relatively unstable due to the
influence of temperature changes and mixing effects. Modeling a relatively conservative
reactive atmospheric species like carbon monoxide has always been viewed as less complex than
modeling a reactive species.  This is analogous to pH and ANC, where ANC is a conserved
species and pH is not.
       To further explain why ANC is emphasized in water quality models is perhaps best
understood by realizing that acidifying atmospheric deposition of nitrogen and sulfur can be
thought as direct inputs of potential acidity (ACY), or stated as negative ANC.  Consequently,
there is well defined linear relationship between potential acidifying deposition and ANC.  This
ANC and deposition relationship facilitates the linkage between ecosystem models that calculate
an ecological indicator and the atmospheric deposition of NOx and SOx.  On the other hand,
there is no direct linear relationship between deposition and pH. There certainly is a
relationship, as acid inputs from deposition lower pH, but the relationship can be extremely
nonlinear and there is no direct connection from a modeling or mass balance perspective between
the amount of deposition entering a system and pH.
       Finally, to illustrate the transient, non-conservative nature of pH, consider two  beakers of
pure distilled water, one closed to the atmosphere and the other open.   The pH of the closed
beaker would be 7, representing neutrality where the concentration of hydrogen ion = the
concentration of hydroxyl ion and their product is 10"14, hence a pH of 7. The beaker open to the
atmosphere receive no inputs of hydrogen ions, but is open to carbon dioxide exchange with the
atmosphere.  Dissolved carbon dioxide turns into carbonic acid, H^COs, and its dissolution ions,
bicarbonate, HCCV, and carbonate, C(V2.  The consequence is that an acid is introduced which
lowers the pH to a value of about 5.7, illustrating the fluctuation of hydrogen ion without change
in hydrogen ion input or export.  This illustration also helps explain the condition of
electroneutrality, as well as explaining why natural aquatic systems often have pH values less
than 7 without anthropogenic inputs and why measuring pH is confounded by the amount of
mixing. Relatedly, rising CC>2 levels are associated with increased acidification and impairment
of oceanic coral reef ecosystems. For now, the effect of rising CC>2 on freshwater systems in the
U.S. is insignificant relative to the strong acid inputs associated with NOx and SOx. Since
carbonate and bicarbonate are negatively charged, the only available positively charged ions to
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counter are hydrogen ion from a limitless supply of water. Essentially, that explains the
definition of an acid which is the ability to affect the transfer of a proton, H+, from water.

       Dissolved Organic Carbon (DOC).  Many humic organic substances produced from
decaying organic matter are considered as weak acids. Natural acidity is often associated with
areas rich in humic susbstances and high levels of DOC.  One of the positive attributes of
defining ANC as the difference of strong cations and anions is that it is an unambiguous
definition compared to the concept of titrating to an equivalence point where all weak anions
become protonated. This is because there are so many different organic acids that would not be
fully protonated at an arbitrary pH of 4.5.  Consequently,  the common use of DOC and ANC to
define pH is not valid for systems with considerable DOC levels as the contributions of weak
organic acid ions also must be balanced by hydrogen ion and therefore pH is lowered relative to
a DOC free system. Many water quality models aggregate all organic acids into a simple
monoprotic (i.e., only one ion representing all  organic acid ions) term with an average
equilibrium constant.  Therefore, knowledge of any three  variables of the pH, ANC, DIG, DOC)
is needed to define the fourth.
       Dissolved Aluminum.  Aluminum species in natural systems is commonly based on the
equilibrium relationships among solid Gibbsite and its dissolved ions (Figure 2-35). Dissolved
aluminum affects the charge relationships in aquatic systems and the distribution of dissolved Al
species is a function of pH.  The prevalence of relatively toxic AL+3  at low pH levels is perhaps
the most direct causative chemical species with regard to  adverse biological effects. From a
chemical characterization perspective, aluminum solubility is treated like other weak acid base
systems, DIG and ANC, and therefore knowledge of 4 of the 5 variables among total dissolved
aluminum, DIG, DOC, ANC and pH uniquely determine the fifth variable.  Collectively, these
principles of mass conservation, equilibrium and electroneutrality are adhered to in water
chemistry models used to estimate key water quality variables as a system responds to
acidification inputs.
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     0
 u
    10
                                  6
                                       pH
8
10
12
 Figure 2-35. Equilibrium diagram of the log of aluminum species concentrations as a function
of pH.  At lower pH there is far greater proportion of dissolved Al species relative to solid
gibbsite (solid aluminum hydroxuide, A1(OH)3) . The most toxics form of Al, free trivalent
aluminum, Al+3 levels rapidly increase with lower pH values.
2.4.2   Bedrock, soil and vegetation processes relevant aquatic acidification.
       The discussion in section 2.4.1 can be thought of as what happens in a lake or stream
watershed relative to inputs from an inflow stream, slope runoff, transport through soils and
direct atmospheric deposition.  Biogeochemical acidification models, discussed below in section
2.4.4, basically attempt to define the chemical makeup of all inputs into a lake that along with
natural chemical conditions and loss processes, determine the chemistry of a surface water.
There are very important processes within a watershed that affect the supply of acids (or strong
anions) and bases (or strong cations to a water body). These include weathering of parent rock
material and soils, which provide a natural supply of base cations, soil cation exchange and
adsorption processes, which influence the quantity, character (chemical species) and delivery of
ions, and vegetation and microbiological processes that modify deposited nitrogen through
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nitrification and uptake. Removal of atmospheric nitrogen, or neutralization, refers broadly to
plant uptake of nitrogen as a nutrient and general immobilization through adsorption in soil and
vegetation layers.
Soil physical/chemical processes and acidification
       Soil  chemical processes lie at the heart of the acidification process. The adsorption and
desorption of anions and cations on soil surfaces, the dissolution of minerals, and the natural
formation of alkalinity in soils control the process of acidification in association with the
deposition and movement of strong  anions (NOs+, SO/f2, Cl") through the soil.
       The  adsorption and desorption of anions and cations on soil surfaces is an important
factor that modifies the effects of acidic deposition to soils. Sulfate is the most important anion
contributing by acidic deposition in most, but not all, parts of the United States. Depending on
the soil characteristics, deposited  S(V2 can move readily through soils into surface water.
However, SO4 2 is less mobile in some areas and is an important factor governing the degree to
which  SO4 2  deposition contributes to soil and water acidification, base cation depletion, and
aluminum (Al) mobilization, each of which can harm biological components of sensitive
ecosystems.
       Sulfur deposition can be adsorbed to soil particles, a process that removes SO/f2 from soil
solution, and therefore prevents leaching of cations and further acidification. The degree to
which  S(V2 adsorbs on soil is dependent on soil characteristics,  in particular the content of clay
minerals.  Soils in the United States that most effectively adsorb SO/f2 occur south of the
maximum extent of glaciation that occurred during the most recent ice age (Rochelle and Church,
1987).  Sulfate adsorption is strongly pH dependent, and a decrease in soil pH resulting from
acidic deposition can enhance the ability of soil to adsorb SO/f2.  The adsorbed SO/f2 acts to
delay the soil and surface water from acidifying. However, this potentially reversible process
results in an accumulation of S in  the soil, which can contribute  to soil acidification if, and when,
that SO/f2 is eventually released back into solution.
       In natural  systems with minimal anthropogenic inputs, an increase in the concentration of
strong-acid anions (NOs+,  SO/f2, Cl") in surface water will be balanced by an equivalent increase
in the concentration of cations.  Thus, neutralization of acidity is controlled by the soil and
involves the release of base cations from the soil into soil water, through weathering, cation
exchange, and mineralization.  Loss of base cations from soil is  a natural process, but the limited
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mobility of anions associated with naturally derived acidity (organic acids and carbonic acid)
controls the rate of base cation leaching under conditions of low atmospheric deposition of S and
N. Because inputs of sulfuric and nitric acid in acidic deposition provide anions that are more
mobile in the soil environment than anions of naturally derived acids (e.g., organic acids and
carbonic acid) these mineral acid anions accelerate natural rates of base cation leaching.
       Soils contain a pool of biological available base cations termed "exchangeable base
cations," which are adsorbed to negatively charged surfaces of soil particles. Base cations can be
released from the soil and enter soil pore water solution by exchanging with other dissolved
cations, such as H+ or Al+3.  Under natural conditions, base cations in the exchangeable pool are
gradually leached from the  soil in drainage water, however, are constantly resupplied through the
weathering of the bedrock and soil. Weathering slowly breaks down rocks and minerals,
releasing base cations to the pool of adsorbed base cations in the soil. The balance between base
cation supply and base cation loss determines whether the pool of available base cations is
increasing or decreasing in  size.  Thus, the main source of cations for acid neutralization in most
watersheds is the accumulated supply of exchangeable base cations in the soil that are mainly
supplied by weathering. Moreover, the size of this supply, and thus the degree to which soil and
surface water acidification occurs, is ultimately determined by the availability of base cations in
watershed bedrock (Webb et al, 1989; Church et al, 1992; Herlihy et al, 1993).
       It has long been  known that leaching of base cations by acidic deposition might deplete
the soil of exchangeable bases faster than they are resupplied, which is termed "base cation
depletion."  Base cation depletion occurs in three-stage process in which buffering of acidity in
the mineral soil is first accomplished by weathering of carbonates and other mineral forms that
weather relatively rapidly.  Once these mineral forms are depleted, buffering is accomplished
largely by cation exchange  on the soil, in which H+ is substituted  for base cations and
concentrations of exchangeable base cations decreases. Once the buffering capacity provided by
cation exchange is depleted, acid neutralization is accomplished by weathering of crystalline
minerals that contain large amounts of silicon (Si) and Al and relatively small amounts of base
cations. At this stage, Al is  mobilized within the soil and exchangeable Al concentrations
increase.
       Therefore,  neutralization of drainage water is accomplished at the expense of soil base
cations. The ability of a soil to exchange base cations between drainage waters is known as the
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cation exchange capacity (CEC) and is determined by many factors.  Soils north of the maximum
extent of glaciation, CEC is largely derived from organic matter, whereas in older southern soils
the surface charge of highly weathered clay minerals is the primary source of CEC. The CEC
derived from organic matter is pH-dependent. Decreases in pH result in a decreases in CEC as
strong-acid anions (NC>3+, S(V2, Cl") are loaded into the soils. The percent base saturation tells
what percent of the exchange sites are occupied by the basic cations. Soil acidification in the
context of acidic deposition can refer to a decrease in soil pH, a decrease in soil percent base
saturation, an increase in Al mobilization, or a combination of these changes.
Conceptualized Model of Acidification
       Galloway  et al. (1983) provided a conceptualized model on how terrestrial systems
undergoing acidification and how base cation concentration in a stream and catchment respond
to a period of elevated inputs  of acidic compounds.  This model can be broken down into 5
stages, starting from the preacidification condition to recovery. Stage I, the period before
acidification, base cations release is equivalent to the rate of chemical weathering plus
atmospheric inputs.  Base cation supply is in steady state equilibrium with cation exchange
surface  and biomass. Stage II, as acid loading increases, the net desorption of cations  increases,
causing base cations to increase in surface waters in order provide  an equivalent countercharge to
the increase in acid anions. Stage III, as acid loading continues soil base saturation is reduced to
a new equilibrium state with respect to acid inputs and the supply of base cations is controlled
only by chemical  weathering, which is relatively unchanged.  The reduced base saturation from
acidification results in decreased export of base cations from desorption, while Al increases.
Stage IV,  as extra acid loading ceases, soils rapidly adsorb cations  released  by chemical
weathering,  thereby reducing  the flux of cations to surface waters and slowing or potentially
halting the recovery of ANC in surface waters associated with decline concentration of acid
anions.  Stage V,  base cations increase as soils reach a new steady  state with chemical
weathering,  atmospheric inputs, and biomass.
       While the  Galloway et al. (1983) model has been fully tested and varied using process-
oriented numerical model (Cosby et al. 1985a,b,c), empirical evidence  for the evolution of these
acidification processes is rare because of the time scale of acidification. Acidification caused by
acidic deposition  takes tens of years or more and there are few time series of appropriate data,
with enough resolution,  and quality data to demonstrate all the processes. The Bear Brooks
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Watershed Manipulation Maine (BBWM), Hubbard Brook Experimental Forest (HBEF) and
other programs has provided empirical evidence of stages II and III and that base cation
deposition in soils has occurred as a consequence of chronic acidification from atmospheric
deposition (Norton et al. 1999). Moreover, many long-term monitoring programs, such as LTM
and TIME with 20 years or more of data, show a slowing of the recovery of ANC in surface
waters despite continued reductions acid anions from acid deposition
(http://www.epa.gov/airmarkets/progress/ARP09_3.html). While this pattern of slowing surface
water recovery follows Galloway's model, it is uncertain what is driving the slowing recovery of
ANC because of limited soil data. Currently, no study has shown an increase in soil base
saturation as a result of decreasing in acid deposition.  Most soil studies continue to show
declining base saturations in areas impacted by acidic deposition (Warby et al 2009).
2.4.3  Biogeochemical ecosystem models used to estimate water quality
       Biogeochemical acidification ecosystem models which incorporate the basic chemical
principles discussed above are important tools to evaluate how multiple environmental factors
alter the relationship between ANC and atmospheric deposition.  Acidification models are
capable of estimating how much acidifying deposition a watershed can accommodate to maintain
a desired ANC, referred to as a critical load (Figure 2-36). The most commonly used models of
acidification are presented in Table 2-8.  These models are designed to be applied at the spatial
scale of the watershed, with the exception of the SMART model.
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Basic approach of steady-state vs. dynamic acidification models
       Acidification models are based on steady-state or dynamic formulations.  The basic
principle of the steady-state approach of aquatic acidification models relates the long term
sustainable ANC to constant levels of acidifying atmospheric deposition.  Because a system
response to time is not required, several simplifying assumptions are invoked which reduces the
model complexity and amount of input data required for execution.  The steady-state models
relate an aquatic ecosystem's critical load to the weathering rate of its drainage basin expressed
in terms of the base cation flux. The weathering of bedrock and soil minerals  is often a major
source of base cation supply to an ecosystem and, therefore, one of the governing factors of
ecosystem critical loads. Dynamic models  include mathematical descriptions of processes that
are important in controlling the chemical response of a catchment. One of the most well-known
dynamic models of aquatic acidification is MAGIC (Cosby et al, 1985a; 1985b;  1985c). It is a
lumped-parameter model of soil and surface water acidification in response to atmospheric
deposition based on process-level information about acidification. "Lumped-parameter" refers to
the extent that spatially distributed physical and chemical processes in the catchment are
averaged or lumped together without affecting the model's reproduction of catchment response.
Process-level information refers to how the model characterizes  acidification  into (1) a section in
which the concentrations of major ions are  assumed to be governed by simultaneous reactions
involving SC>42" adsorption,  cation exchange, dissolution-precipitation- speciation of aluminum,
and dissolution-speciation of inorganic carbon; and (2) a mass balance section in which the flux
of major ions to and from the soil is assumed to be controlled by atmospheric inputs, chemical
weathering, net uptake and loss in biomass  and losses to runoff.  At the heart of MAGIC is the
size of the pool of exchangeable base cations in the soil.  As the fluxes to and from this pool
change over time owing to changes in atmospheric deposition, the chemical equilibria between
soil and soil solution shift to give changes in surface water chemistry. The degree and rate of
change of surface water acidity thus depend both on flux factors and the inherent characteristics
of the affected soils.
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Table 2-8.  Summary  several commonly used  acidification models (See ISA Annex A for
more comprehensive list and discussion of acidification models).
Model name
Dynamic
or steady
state
Model description
Steady-state mass
balance models
/Steady-state water
Chemistry
(SSWC)/
Steady-
state
The basic principle is based on identifying the long-term
average sources of acidity and alkalinity in order to determine
the maximum acid input that will balance the system at a
biogeochemical safe-limit. Several assumptions have been
made in the steady state calculations. First, it is assumed that
ion exchange is at steady state and there is no net change in
base saturation or no net transfer of ANC from soil solution to
the ion exchange matrix. It is assumed that for N there is no net
denitrification, adsorption or desorption and the N cycle is at
steady state. Sulfate is also assumed to be at steady state: no
sulfide oxidation, sulfate uptake, sulfate permanent fixation or
sulfate reduction are significant. Simple hydrology is assumed
where there is straight infiltration through the soil profile.	
First-order Acid
Balance model
(FAB)
Steady-
state
The FAB model includes more explicit modeling of N
processes including soil immobilization, denitrification, and
wood removal, in-lake retention of N and S, as well as lake
size.
Model of
Acidification of
Groundwater in
Catchment
(MAGIC)
Dynamic
MAGIC is a lumped-parameter model of intermediate
complexity, developed to predict the long term effects of acidic
deposition on surface water chemistry. The model simulates
soil solution chemistry and surface water chemistry to predict
the monthly and annual average concentrations of the major
ions in these waters. MAGIC consists of: a section in which the
concentrations of major ions
are assumed to be governed by simultaneous reactions
involving SO4 2- adsorption, cation exchange, dissolution-
precipitation- speciation of aluminum, and dissolution-
speciation of inorganic carbon; and a
mass balance section in which the flux of major ions to and
from the soil is assumed to be controlled by atmospheric inputs,
chemical weathering, net uptake and loss in biomass and losses
to runoff.
PnET-BGC
Dynamic
PnET/BGC simulates major biogeochemical processes, such as
forest canopy element transformations, hydrology, soil organic
matter dynamics, N cycling, geochemical weathering, and
chemical equilibrium reactions in solid and solution phases, and
allows for simulations of land disturbance. The model uses
mass transfer relationships to describe weathering, canopy
interactions and surface water processes. Chemical equilibrium
relationships describe anion adsorption, cation exchange and
soil solution and surface water speciation. The model can be set
to operate on any time set, but is generally run on a monthly
time-step. It is applied at the stand to small-watershed scale.
DayCent-Chem
Dynamic
DayCent-Chem links two widely accepted and tested models,
one of daily biogeochemistry for forest, grassland, cropland,
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Model name
Dynamic
or steady
state
Model description
                               and savanna systems, DayCent (Parton et al., 1998), and the
                               other of soil and water geochemical equilibrium, PHREEQC
                               (Parkhurst and Appelo, 1999). The linked DayCent/PHREEQC
                               model was created to capture the biogeochemical responses to
                               atmospheric deposition and to explicitly consider those
                               biogeochemical influences on soil and surface water chemistry.
                               The linked model expands on DayCent's ability to simulate N,
                               P, S, and C ecosystem dynamics by incorporating the reactions
                               of many other chemical species in surface water.	
Very Simple
Dynamic (VSD)
soil acidification
model
Dynamic
This model is frequently used in Europe to simulate
acidification effects in soils when observed data are sparse. The
VSD model consists of a set of mass balance equations,
describing the soil input-output relationships, and a set of
equations describing the rate-limited and equilibrium soil
processes. It only includes weathering, cation exchange, N
immobilization processes, and a mass  balance for cations,
sulfur and N. In the VSD model, the various  ecosystem
processes have been limited to a few key processes. Processes
that are not taken into account include canopy interactions;
nutrient cycling processes;  N fixation and NH4 adsorption;
SO42- transformations (adsorption, uptake, immobilization,
and reduction); formation and protonation of organic anions;
and complexation of Al.	
Simulation Model
for Acidification's
Regional Trends
(SMART)
Dynamic
The the SMART model consists of a set of mass balance
equations, describing soil input/output relationships, and a set
of equations describing the rate-limited and equilibrium soil
processes. It includes most of the assumptions and
simplifications given for the VSD model. SMART models the
exchange of Al, H, and divalent base cations using Gaines
Thomas equations. Additionally, SO4
2- adsorption is modeled using a Langmuir equation (as in
MAGIC) and organic acids can be described as mono-, di-, or
tri-protic. The SMART model has been developed with regional
applications in mind, and an early example of an application to
Europe can be found in De Vries et al. (1994).	
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    Current Condition of Acidity
             and Sensitivity
   Criticial Load
     meq/m2/yr
    • Highly Sensitive: < SO
      Moderately Sensitive: 51 -100
      Low Sensitivity: 101 - 200
    • Not Sensitive: > 201
   |  | Adirondack Boundary
Source: EPA 2009
                  Current Condition of Acidity
                          and Sensitivity
                  Criticial Load
                    meq/m2/yr
                   • Highly Sensitive: < 50
                     Moderately Sensitive: 51-100
                     Low Sensitivity: 101 - 200
                   • Not Sensitive: > 201
                                                 Source: EPA 2009
Figure 2-36.  Critical loads of acidifying deposition based on MAGIC modeling  that
each surface water location can receive in the Adirondack and Shenandoah Case Study
Areas while maintaining or exceeding an ANC concentration of 50 ueq/L based on 2002
data. Watersheds with critical load values <100 meq/m2/yr (red and orange circles) are
most sensitive to surface water acidification, whereas watersheds with values >100
meq/m2/yr (yellow and green circles) are less sensitive sites.
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Trajectory of recovery for ecosystems from CL calculated by steady-state vs. dynamic
acidification models
       Steady-state models assume that the ecosystem is in equilibrium with the critical load of
deposition; therefore the long-term sustainable deposition is indicated. This is the relevant
information needed to provide protection from deposition in perpetuity as the system comes into
equilibrium with the pollutant critical load (ISA Appendix D). In the U.S., few (if any)
ecosystems qualify as steady-state systems.  Therefore the assumption of equilibrium in the
steady-state model is often false. This has implications for the temporal aspects of ecosystem
recovery.  The steady-state models give no information concerning the time to achieve the
equilibrium or what may happen to the receptor along the path to equilibrium.  The recovery of
an ecosystem based on a critical load from a steady state model may take several hundred years.
In other words the assumption that attainment of deposition values below the steady-state critical
load will result in biological recovery within a specified time period may not be valid.
       Dynamic models calculate time-dependent critical loads and therefore do not assume an
ecosystem is in equilibrium. This is the relevant information needed to provide protection from
damage by the pollutant within a specified time frame. As a general rule, the shorter the time
frame selected, the lower the critical load.
       The most comprehensive study done in the United States is Holdren et al.  1992 that
compared critical loads calculated by the dynamic MAGIC model versus SSWC steady-state
approach.  A 50-yr simulation critical load was obtained from the MAGIC model. Holdren et al.
1992 found that both models yielded the same general trends.  The critical load estimates
projected using the dynamic versus steady-state models are  consistently higher. Both models
produced critical load values approximately equal for systems with critical loads of about zero.
However, at higher critical load values the two model outputs diverge rapidly, implying that
watersheds with larger inherent buffering capacities respond more slowly to a given level of
acidic deposition.  The apparent reason for this is that the watersheds represented by the
dynamic model retain a larger fraction of their buffering capacity in the base cation exchange
pool for the 50-year time  scale of the simulation.  In the steady-state models, the cation
exchange pool is assumed to be in equilibrium and does not provide additional buffering.
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Data Requirements of steady-state vs. dynamic acidification models
       There are various factors that modify the ANC to deposition relationship, which are
described by models that parameterize ecosystems to simulate the process of acidification.  The
steady-state models used for critical loads analysis in the REA required input data for between
17 and 20 variables, including water chemistry data from the TIME and LTM programs, which
are part of the Environmental Monitoring and Assessment Program (EMAP). A summary of the
variables for steady state models (and data sources for the calculations made in the REA) is
given in Appendix A.
        The data requirements required to run dynamic models, such as MAGIC, are greater.
The equations that characterize the chemical composition of soil water in MAGIC contain 33
variables and 21 parameters (Cosby et al. 1985a). Data required to conduct dynamic modeling
are not available for as many places as the data required to conduct steady-state modeling.
Comparison of two steady-state models: FAB and SSWC
       The steady state models used in the REA were the Steady State Water Chemistry model
(SSWC), and the First-order Acid Balance model (FAB).  The SSWC and FAB  models were
used to calculate critical loads for specified ANC levels in the case study areas.
       The SSWC and FAB make different assumptions of ecosystem function.  Most notably,
biogeochemical pathways of N deposition are considered differently in the two models. In the
SSWC model, sulfate is  assumed to be a mobile anion (i.e. S  leaching = S deposition), while
nitrogen is retained in the catchment by  various processes.  The assumption that  all N is retained
by the ecosystem and does not contribute to acidification is incorrect because in many
ecosystems nitrate leaching is observed. If nitrate is leaching  out of an ecosystem, it cannot also
be true that it has all been retained. Nitrate leaching is determined from the sum of the measured
concentrations of nitrate in the runoff. The critical  load for sulfur that is calculated by SSWC
can be corrected for the amount of nitrogen that contributes to acidification. When an
exceedance value for the critical load is  calculated, the critical load is subtracted from S
deposition plus the amount of nitrate leaching, as it represents the difference between N
deposition and N retention by  the ecosystem.  N leaching data used in this calculation are
considered robust.
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       In contrast to the SSWC approach, the FAB model includes more explicit modeling of N
processes including soil immobilization, denitrification, and wood removal, in-lake retention of
N and S, as well as lake size. Although N cycling is more detailed in the FAB model, there is
greater uncertainty in the input data needed to characterize the components of the N cycle. The
FAB model yields a deposition load function for a specified level of an endpoint. This function is
characterized by three nodes that are illustrated on Figure 2-37, 1. the maximum of amount of N
deposition when S deposition equals zero (DLmax (N)); 2. the amount of N deposition that will
be captured by the ecosystem before it leaches (DLmin(N)); and 3. the maximum amount of S
deposition considering the N captured by the ecosystem (DLmax (S)).  The function represents
many unique pairs of N and S  deposition that will equal the critical load for acidifying
deposition.
       The three models, MAGIC, SSWC and FAB, discussed above were widely used in the
REA and this PA. MAGIC  enabled the construction of time series estimates of water quality
change and provided a more scientifically rigorous model to conduct comparisons with certain
parameterizations used in steady state models.  As will be described in Section 7, attributes of
SSWC and FAB modeling,  several of which are illustrated in Figure 2-37, were incorporated in
constructing the form of the standard.
                                                    H Deposition
Figure 2-37. Illustration of a generalized N + S deposition tradeoff curve that is calculated by
using the FAB approach.
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2.4.4   Water Quality Networks
       An overview of surface water chemistry monitoring programs is incorporated in EPA's
ISA (EPA, 2008) and is the basis for much of this discussion.  The TIME/LTM program is
described immediately below.  A summary of the water quality data bases used in the PA are
provided in Table 2- 9. National alkalinity and ANC maps, which help support development of
acid sensitive and less sensitive categories which are considered in developing the standard (see
section 7.2.5), are included at the end of this sub-section (Figures 2-39 and  2-40). Appendix D
provides data summaries of water chemistry variables (SO/t, natural base cation supply, DOC,
ANC) delineated by Omernik Ecoregion Level III categories.
TIME/LTM Program descriptions
       The Temporally Integrated Monitoring of Ecosystems (TIME) and the long-Term
Monitoring (LTM) programs (Table 2-10; Figure 2-38) are complementary EPA surface water
monitoring networks that provide information on a variety of indicators necessary for tracking
temporal and spatial trends in environmental response to changes in regional air quality and acid
deposition in ecosystems sensitive to acid rain in the eastern United States.  Some of these
indicators include, but are not limited to:  acid neutralizing capacity (ANC),  sulfate (SC>42"), and
nitrate (N(V).  Both programs are operated cooperatively with numerous collaborators in state
agencies, academic institutions, and other federal agencies.
       The TIME program was developed as a special study within EPA's Environmental
Monitoring and Assessment program (EMAP) to track, in more detail, the trends in acid relevant
chemistry of particular classes of acid sensitive lakes and streams in the eastern United States.
TIME lakes, located in the Adirondacks and New England, have been sampled annually each
summer since  1991, while TIME streams, located in the Ridge and Blue Ridge Provinces and the
Northern Appalachian Plateau,  have been sampled annually in spring since 1993. Based on the
concept of a probability sample, TIME sites were statistically selected to be representative of a
larger, target population, thus results from TIME samples  can be extrapolated, with known
confidence, to the target populations as a  whole.
       The LTM program focuses on detecting long-term trends in acid relevant chemistry in
lakes and streams across a gradient of acidic deposition. LTM sites are a subset of sensitive
lakes and streams in the eastern United States with long-term data that,  in most cases, date back
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to the early 1980s.  Sites are sampled 3 to 15 times per year to provide information on how the
most sensitive of aquatic systems in each region are responding to changing deposition, as well
as give information on seasonal chemistry and episodic acidification. LTM lake sites are located
in the Adirondacks and New England while LTM stream sites are located in the
Catskills/Northern Appalachian Plateau and the Ridge and Blue Ridge Provinces.

Table 2-9.  Summary of data sources considered for the evaluation of national ANC.
Program
EPA Long Term Monitoring Vermont (LTM VT)
EPA Eastern Lakes Survey (ELS)
Adirondack Lake Survey (ALS)
EPA Western Lake Survey (WLS)
EPA National Stream Survey (NSS)
VTSSS
EPA Long Term Monitoring Colorado sites
(LTM CO)
EPALong Term Monitoring Midwest Sites
(LTM MW)
VT SSS LTM
EPA Long Term Monitoring Pennsylvania
sites(LTM PA)
EPA Long Term Monitoring Catskill sites
(LTM CAT)
EPA Long Term Monitoring: Annual average from
1992-2007
EPA EMAP Northeast Lake Survey
EPA Long Term Monitoring Maine sites
(LTM ME)
Regional Environmental Monitoring
Program Maine sites (REMAP ME)
EPA EMAP Mid Atlantic streams (EPA EMAP
MAIA)
EPA EMAP Mid Atlantic streams (EPA EMAP
MAIA)
EPA EMAP Western Stream and River Survey
(EMAP WEST )
EPA National Lakes Survey (NLS)
USGS NAWQA Program
EPA Storet Program
Dates of
observations
1983-2007
1984
1984-1987
1985
1986
1987 & 2000
1990-1994
1990-2000
1990-2007
1990-2007
1990-2007
1990-2007
1991-1994
1992-2007
1993
1993-1996
1997-1998
2000-2004
2010


Reference
EPA/903/R-00/015
EPA/620/R-93/009
Stoddard.et.al.WRR.1996
EPA 620-R-05-005
Stoddard.et.al.WRR.1996
EPA841-B-06-002
Stoddard.et.al.WRR.1996
Stoddard.et.al.WRR. 1996
Stoddard.et.al.WRR.1996
Stoddard.et.al.WRR. 1996
EPA905-R-92-001
EPA/600/4-88/032
Stoddard.et.al.WRR. 1996
regl qa.pdf
Stoddard.et.al.WRR. 1996
EPA/R-06/XX
EPA-600-388-021a
EPA/600/3-86/054b
EPA841-F-09-007.
http://water.usgs.gov/naw
qa/
http ://www. epa.gov/store
t/
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                          Adirondack Mountain
                     Catskills/Northern Appalachian Plateau
                        Ridge and Blue Ridge Provinces
    LTM Lakes
•   LTM Streams
•   TIME Lakes
°   TIME Streams
Figure 2-38. Active TIME/LTM sampling locations.
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  Table 2-9.  Characteristics of TIME/LTM Sites
Sites
TIME Lakes**
Adirondack^ (New York)
Maine
- Massachusetts
- Maine
- New Hampshire
- Rhode Island
- Vermont
Total TIME Lakes
TIME Streams**
Northern Appalachians
- Pennsylvania
- West Virginia
Ridge /Blue Ridge
- Maryland
- Pennsylvania
- Virginia
- West Virginia
Total TIME Streams
Total TIME Lakes and
Streams
LTM Lakes
Adirondack^ (New York)
Maine
Vermont
Total LTM Lakes
LTM Streams
Appalachians
- Pennsylvania*
Catskills
-New York ±
Virginia Intensive ±
Virginia Extensive (Trout
Streams)
Total LTM Streams
Total LTM Lakes and
Streams
TOTAL TIME/LTM
SITES

43
8
5
14
1
1
72

21
14
1
3
13
4
56
128


52
16
12
SO

5
4
3
64
76
156

284

Collection Major ions Total AT Al .
interval* collected Collected Speciated „ . ,.
Speciation

Summer/Fall
Summer
Summer
Summer
Summer
Summer


Spring
Spring
Spring
Spring
Spring
Spring




Monthly
Quarterly
Quarterly


Monthly
Monthly/Episodes
Weekly/Episodes
Quarterly






X XX"
X XX"





X XX"
X XX"




X XX"
X X " X
X XX"


X X " X
X XX"
X " " X
X " " X





* Samples are collected once per specified season/interval
** All TIME sites are monitored annually
± Stream flow data are collected from these sites
                                                        2-81

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                                 Total Alkalinity of Surface Waters
Figure 2-39.  Surface water alkalinity in the conterminous U. S.  Shading indicates the range of alkalinity within
which the mean annual values of most of the surface waters of the area fall (Omernik and Powers), 1983.
  Degree of Acidification in Acid
  Sensitive Lakes and Streams
        o  Acute Acidification     < Q
        <"  SevereAciclification    g-20
        •  Elevated Acidification   20-50
        »  Moderate Acidification   50-100
        *  LowAcidification      .^QQ
 Figure 2-40. National map of ANC data (ueq/L) based on historical and contemporary data sorted by ANC classes.
                                                      2-82

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2.5    TRENDS
       This section provides summaries of time series trends of emissions, air quality, deposition
and water quality relevant to oxides of nitrogen and sulfur and aquatic acidification. The value
of this summary is in bringing the four components in the source-to-effects continuum (Figure 2-
41) together in one place to provide general trends information but also to develop, at least
conceptually,  a retrospective view of how the basic components which underlie the form of the
standard (section 7.2) respond together.  Extensive use is made of EPA's own annual reports on
progress associated with the Acid Rain Program (EPA, 2010). This information is a response to
suggestions that the form of the standard be studied in a  "hindcast" manner. Unfortunately,
historical data and modeling results are not available to adequately support calculation of the
standard as described in chapter 7. Consequently, these  combined time series examples are used
to demonstrate in an associative manner that the basic tenets of the conceptual model of the
standard are valid for the standard being developed in this assessment, which relies on similar
directional changes in the source to effects continuum.  One could identify the acid rain
emissions limits as part of Title IV as a marker to assess progress that is specific to the mission
component at the beginning of the source-to-effects continuum.  Similarly, the standard
developed  in this assessment might be thought of as providing an atmospheric marker upon
which future progress can be judged.
        Secondary and deposition loads
        Visibility, ac id if icat ion,
        eutrophication, metals
      Exposures
      Inhalation, digestion
                 Ecosystem * effects
                 Defoliation, Visibility
                  biodiversity,
                 Metals concentration
                                     Increasing influence in
                                     confounding factors and
                                     perceived value to public policy
                                               Feedback/correction
Figure 2-41.  Conceptual source to effects pipeline diagram illustrating basic accountability concepts as one
proceeds from source emissions through the air and eventually to effects.
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2.5.1 Emissions
       Sickles and Shadwick (2007) summarize NOx and SOx emission changes from 1990 -
2004. Title IV (Acid Rain provisions) of the Clean Air Act Amendments of 1990 (CAAA)
established phased year-round controls for SO2 and NOx emissions from electrical generating
units (EGUs) that became effective in!995 for SC>2 and 1996 for NOx (Phase I); while additional
controls became effective under Phase II in 2000.  Over half of the EGUs targeted by the CAAA
are in six states located along the Ohio River (IL, IN, KY, OH, PA, and WV). Beginning in
1999, O3 season (i.e., summer) NOx controls focusing on EGUs became effective in selected
eastern states under the Ozone Transport Commission (OTC) and were  superseded in 2003 by
the NOx State Implementation Plan (SIP)  Call. The affected states have chosen to meet their
mandatory SIP Call NOx reductions by participating in the NOx Budget Trading Program, a
market-based cap and trade program for EGUs and large industrial units. In addition, various
mobile source NOx emissions control programs began in the mid to late 1990s. These mobile
source programs have a cumulative effect  of reducing NOx emissions over time as the mobile
fleet is replaced.
       The recent declines from 2005 to 2009 in NOx and  SOx emissions (Figures 2-42 and 2-
43) are attributed to continued implementation of NOx SIP CALL and transportation sector
rules, as well efforts to implement controls associated with the Clean Air Interstate Rule (CAIR).
Annual reports explaining the rules and programs  addressing EGUs and the relationships
between emission reductions and air quality and deposition changes is provided by EPA's Clean
Air Market Division at http://www.epa.gov/airmarkets/progress/progress-reports.htm.  Major
reductions in EGU SOx and NOx emissions are associated  with market trading of emissions
which was a tool in implementing emissions reduction targets in Title IV and other rules.
                                         2-84

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    35
    30
  o 25
  ; 20
    15
    10
                                                  -*- NOx  -»- S02
Figure 2-42 Time series trends of all anthropogenic NOx and SOx emissions based on EPA's National Emissions
Inventory (NEI).
        1990   1995    1996    1997   1998    1999   2000   2001    2002   2003    2004   2005    2006   2007
                      • NOX Program Affected Sources     | Title IV Sources Not Affected for NO,
                                                                                                 2008    2009
         1980   1985   1990   1995   1996   1997   1998   1999   2000   2001   2002   2003   2004  2005   2006   2007   2008   2009
           • Phase I (1995-1999) Sources    • Phase II (2000 on) Sources     • All Affected Sources     • Allowances Allocated
Figure 2-43. Trends of NOx (above) and SOx (below) reductions associated with EPA's Acid Rain Program
(ARP) which established market trade units for NOx and SOx emissions associated with EGUs (EPA, 2010).
Trends of SO2 reductions associated with EPA's Acid Rain Program (ARP) which established market trade units for
NOx and SOx emissions associated with EGUs. ARP units have reduced annual SO2 emissions by 67 percent
compared with 1980 levels and 64 percent compared with 1990 levels  (EPA, 2010).
                                                   2-85

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2.5.2  Air quality and deposition
       Significant reductions in ambient air concentrations in the Eastern U.S. of 862, NOs and
864 over the last two decades (Figures 2-44 and 2-45) are consistent with emission trends for
NOx and SOx over the same period.  Trends of observed wet deposition have been provided as a
feature product by the NADP; sulfate and nitrate wet deposition patterns (Figure 2-46) generally
exhibit the same features of the analogous air quality patterns.  Dry deposition patterns can be
assumed to exhibit identical patterns of air quality concentrations, given the direct dependence of
dry deposition on concentration. Sickles and  Chadwick (2007) provide quantitative  estimates
linking changes in emissions, concentrations and deposition patterns from 1990 to 2004.  Their
findings (Table 2-11) broken down by pollutant and three subregions across the Eastern U.S.
demonstrate similar patterns of reductions through emission, concentrations and deposition.
However, stronger correlations between emissions changes and total oxidized nitrogen and sulfur
in concentration and deposition fields relative to relationships between emissions and a single
species.
Table 2-11 Period 1 Emissions Density and Period 1-to-Period 3 Relative Changes (%) in Oxidized Sulfur and
Nitrogen Emissions, Atmospheric Concentration, and Dry, Wet, and Total Deposition (from Sickles and Chadwick,
2007).

Region
E
MW
NE
SO

Region
E
MW
NE
SO
Emissions
PI Density1
37
50
38
28
Emissions
PI Density"
19
20
25
15

SO,, °i,
-39
^49
-38
	 30

NOx. %
22
-21
-35
-14
>\tnii
SO-,. "„
	 33
	 34
-31
	 35
Atnn
HNO.s, %
- 1 3
	 1 1
	 1 8
-10
iisphcnc Concentration
so; , "„ S, ",,
-2 1 -30
-22 -32
-21 -29
	 20 	 30
i>sphcnc Concentration
NO., , % OxN, "-.
11 4
8 -1
19 -8
Q ^5

Dry S. %
	 3 1
	 29
-30
	 33

Drv OxN, "-«,
8
3
P
8
^^Deposilion^
Wet S, "-„
	 22
	 27
-20
	 19
Deposition
Wet OxN, °«
-12
	 1 2
	 14
- 1 0

Total S. %
-26
-28
-25
	 25

Total OxN. ".«
1 1
-8
_ 14
_9
  "'SO^ emissions density in units of kgS/'(ha y).
  ^NOx emissions density in units of kgN/( ha y ).
       Trends in reduced nitrogen are based only on ammonium ion observations, given the very
limited availability of ammonia gas monitoring.  Patterns of reduced nitrogen (Figure 2-47)
based on ammonium ion are difficult to interpret with respect to trends of precursor ammonia
emissions, which are not presented in this section.  Ambient ammonium levels have decreased
while wet deposition has increased over the same period. Because ammonium is associated
atmospheric nitrate or sulfate, reductions in NOx and SOx emissions that lead to reductions in
                                           2-86

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atmospheric nitrate and sulfate also lead to reduced atmospheric ammonium.  However, because
ammonia gas is a precursor for ammonium, and ammonium is dependent on sulfate and nitrate,
very little can be inferred regard from atmospheric ammonium levels in regard to total
contribution to reduced nitrogen.  One can infer that dry deposition of ammonium has been
reduced, as has the contribution of ammonium to particulate matter mass.  The sum of dry and
wet ammonium deposition is a better indicator relative for associating atmospheric data with
trends of ammonia emissions. While dry deposition trends are not presented here, wet
ammonium levels have increased and it is reasonable that infer that total reduced nitrogen
(NHx), which is associated with ammonia emissions, probably has not changed in the same
manner as NOx and SOx emissions.
       Kim et al (2006) quantified changes in NOx emission reductions and NO2 concentrations
for New York and Ohio, using satellite column NO2 observations to capture region wide NO2
patterns (Figure 2-48).  The use of satellite data to capture region wide NO2 patterns is indicative
of the shortage of available NO2 data in rural locations, as described earlier in section 2.2.
       Before describing water quality trends, it is informative to note changes in pH over the
last two decades. In this context, pH serves simply as an indicator and does reflect that
concurrent reductions in emissions of acid generating gases, NOx and SOx, clearly have resulted
in significant reductions in rain water hydrogen ion levels (Figure 2-49).   The pH patterns
continue to reflect a sulfur dominated air regime in the Eastern U.S., relative to the rest of the
contiguous U.S.
                                          2-87

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Figure 2-44.  Eastern U.S. annual average spatial distribution of SO2 (left) and total nitrate (right) concentrations
averaged over 1989 -1991 (top) and 2007 - 2009 (bottom).  Data are based on EPA's CASTNET program.
                                                 2-88

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                                                    Souire: CASTNET
                                                                                        I '.I I .  .Mil	1
 Figure 2-45. Eastern U.S. annual average spatial distribution of SO4 (left) concentrations averaged over 1989 -
 1991 (left) and 2007 - 2009 (right). Data are based on EPA's CASTNET program.
                1989-1991
                                                                             1989-1991
 Figure 2-46. U.S. annual average spatial distribution of wet sulfate (left) and nitrate (right) deposition averaged
over 1989-1991 (top) and 2007 - 2009 (bottom) based on the NADP.
                                                2-89

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                                         I ',] !' '., i V.1IJU . •' "'i
                                                         Source: CASTOET
                                                          Source, CASTNET
Figure 2-47. U.S. annual average spatial distribution of wet ammonium sulfate deposition (left) and ambient air
ammonium concentrations (right) averaged over 1989 -1991 (top) and 2007 - 2009 (bottom) based on CASTNET
and NADP.
             • Ovo : Satellite
              Ohio : Emission inventory
             • New York : Satellite
              New York . Emission inventory
                   i
                         i
                              i
A VC NO2
[molec crrv2 yr
  I >
   6.0 1014

   4.0 10"

   2.010"

   0.0 10°°

  J-2.0 10"

  1-4.0 10"
                                                        D  -90   -60
                                                                   -30    0   30
                                                                      Longitude
                                                                                 60   90   120  150
Figure 2-48. Left - superimposed Eastern U.S. emission and combined GOME and SCIAMACHY NO2 1997-2002
trends (Kim et al., 2006); right - GOME NO2 trends from 1995 - 2002 (after Richter et al., 2005). Clear evidence of
reductions in midwest U.S. and European NOx emissions, and increased NOx generated in Eastern Asia
                                                    2-90

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                                            '.' " . .Mil"  	
   	 ••,,11)1
                                           USEPA/CAMO O7«to-I0
Figure 2-49. U.S. annual average spatial distribution of hydrogen ion concentration in rain water as pH  averaged
over 1989-1991 (top) and 2007 - 2009 (bottom) based on the NADP.
2.5.3  Water quality
       General patterns in the trends of major anions directionally track emission changes over
the last two decades (EPA, 2010; Figures 2-50 to 2-52).  Summaries of water quality trends prior
to 2000 (Driscoll et al, 2003;  Stoddard, 2003; Figures 2-51 to 2-52) illustrate the initial period of
declining SOx emissions with evidence of decreasing anions and increasing ANC.  Several
                                             2-91

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studies have illustrated the general trend of decreasing water column levels of strong ions and
increasing ANC with reductions in SOx and NOx deposition are synthesized in EPA's ISA.
       However, as one proceeds through the source to effects continuum, strong direct
relationships gradually diminish at each step as there are a myriad of confounding factors that
start to affect each stage. This is particularly true in moving from deposition to water quality,
where a variety of ecosystem processes moderate nitrogen and sulfur deposition both in terms of
chemical transformations and delivery rates associated with soils and vegetation processes
discussed above.  For example, there are  directionally different responses of water column
sulfate (e.g., increase in southern Appalachians, decreases in Adirondacks) to declining levels of
sulfate deposition (Figure 2-50).  The inherent lag in recovering from acidification largely
associated with soil adsorption and exchange processes implies that several decades of
information may be required to sort out long term responses in water column chemistry relative
to changes in emissions, air and deposition but the relative degree of responsiveness will be
influenced strongly by watershed soil and vegetation characteristics.  A recognition of the
inherent lag in ecosystem response to changes in atmospheric variables is necessary to
conceptualize the linking of air quality to  water quality that rely on steady state models
ecosystem models which provide the long term sustained response of water chemistry to
atmospheric conditions. This understanding underlies the basic difference between the air
quality water indices of acidification, both of which are incorporated in the form of the standard
(section 7.2). An air quality acidification  index reflects the eventual steady state conditions in
aquatic systems that would be achieved assuming atmospheric a given atmospheric state of air
quality conditions. The time lag in those conditions between aquatic and atmospheric media can
range from near real time to decadal differences.
       For example, the analyses of the Adirondack and Shenandoah  Case Study Areas
indicated that although wet deposition rates for 862 and NOx have been reduced since the mid-
1990s,  current concentrations are still well above simulated pre-acidification (1860) conditions
(Figures 2-53 to 2-56).  Modeling predicts NCV and SC>42" are 17- and 5-fold higher,
respectively, in 2006 than under simulated pre-acidification conditions. Based on the 2006
Model  of Acidification of Groundwater in Catchment (MAGIC) simulations, the estimated
average ANC across the 44 lakes in the Adirondack Case Study Area is 62.1  ueq/L (± 15.7
                                           2-92

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ueq/L); 78 % of all monitored lakes in the Adirondack Case Study Area have a current risk of
Elevated, Severe, or Acute.  Of the 78%,  18% are chronically acidic (REA 4.2.4.2).
                                           1990-2008 Acid Neutralizing
                                           Capacity (ANC)
                                              Increasing significant trend
                                              Increasing non-significant trend
                                           O Decreasing non-significant trend
                                           • Decreasing significant trend
                                        1990-2008 Sulfate Ion Concentration
                                        • Increasing significant trend
                                        O Increasing non-significant trend
                                        o Decreasing non-significant trend
                                        • Decreasing significant trend
Figure 2-50.  Generalized trends in water quality variables ANC (above) and sulfate ion (below) - EPA, 2010.
                                                 2-93

-------
S04-
N03-
N03-

-------
o
                140
                120
            8?
            O 3.
            O"'
            O
            z
      80
      60
      40
      20
                   1850
                    1900
1950
2000
2050
Figure 2-53.     Average NO3" concentrations (orange), SO42~ concentrations (red), and ANC (blue) across
the 44 lakes in the Adirondack Case Study Area modeled using MAGIC for the period 1850 to 2050.
             ANC Preacidification (1860) and Current Condition (2006)
               Preacidification (1860)
                                        Current (2006)
        Source: EPA 2009
                                  ANC (|Jeq/L)
                                    •  <0
                                       0-20
                                       20-50
                                       50-100
                                    •  >100
Figure 2-54.     ANC concentrations of preacidification (1860) and 2006 conditions based on hindcasts of
               44 lakes in the Adirondack Case Study Area modeled using MAGIC.
                                             2-95

-------
0
To
c
u
c
o
U
U


100
_ 80-
j"
o- 60 -
^i

4u

20
0
            1850
Figure 2-55.     Average NO3" concentrations orange), SO4 "concentrations (red), and ANC (blue) levels
                for the 60 streams in the Shenandoah Case Study Area modeled using MAGIC for the

g
— i
•Hj
aJ
c
o
u
-z.

120 -,
100 -
_ 80 -
o- 60 -
4)
-40
20-
0 -
            1850
2050
Figure 2-56. ANC levels of 1860 (preacidification) and 2006 (current) conditions based on hindcasts of 60 streams
in the Shenandoah Case Study Area modeled using MAGIC
                                                2-96

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2.6 REFERENCES

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     EPA/600/R-92/186). Corvallis, OR: U.S. Environmental Protection Agency.
Cosby BJ; Hornberger GM; Galloway JN. 1985a. Modeling the effects of acid deposition:
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Cosby BJ; Wright RF; Hornberger GM; Galloway JN. 1985b. Modelling the effects of acid
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Cosby BJ; Hornberger GM; Galloway JN; Wright RF. 1985c. Time scales of catchment
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Dennis, R., R. Mathur, J.E. Pleim and J.T. Walker. 2010. Fate of Ammonia Emissions at the
     Local to Regional Scale as Simulated by the Community Multiscale Air Quality Model,
     Atmospheric Pollution Research.
de Vries W; Reinds GJ; Posch M; Kamari J. 1994. Simulation of soil response to acidic
     deposition scenarios in Europe. Water Air Soil Pollut, 78, 215-246.
Driscoll CT; Driscoll KM; Roy KM; Mitchell MJ. 2003 c. Chemical response of lakes in the
     Adirondack Region of New York to declines in acidic deposition. Environ. Sci. Technol.,
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Finlayson-Pitts BJ;  Pitts JN Jr. 2000. Chemistry of the upper and lower atmosphere: theory,
     experiments and applications. San Diego, CA: Academic Press.
Galloway JN; Norton SA; Church MR. 1983. Freshwater acidification from atmospheric
     deposition of sulfuric acid: A conceptual model. Environ Sci Technol, 17, 541A-545A.
Galloway, J. N., G. E. Likens, and M. E. Hawley. 1984. Acid precipitation: natural versus
     anthropogenic components. Science 226:829-831.
Grantz D.A., J.H.B. Gardner and D.W. Johnson.  2003. Ecological effects of particulate matter,
     Environment  International, 29 (2003) 213-239.
Hemond. 1990. Acid Neutralizing Capacity, Alkalinity, and Acid-Base  Status of Natural Waters
     Containing Organic Acids, Environ Sci. Technol, 24, 1486-1890.
Herlihy, A. T., P. R. Kaufmann, M. R. Church, P. J. Wigington, Jr., J. R. Webb, and M. J. Sale.
     Rochelle, B. P., and M. R. Church. 1987. Regional patterns of sulfur retention in
      watersheds of the eastern U.S. Water Air Soil Pollut. 36:61-73.
Herlihy AT; Kaufmann PR; Church MR; Wigington PJ Jr; Webb JR; Sale MJ. 1993. The effects
      of acid deposition on streams in the Appalachian Mountain and Piedmont region of the
      mid-Atlantic United States. Water Resour. Res. 29:2687-2703.
Holdren, G.R., T.C. Strickland, P.W.Shaffer, P.F. Ryan, P.L. Ringold and R.S. Turner. 1992.
     Sensitivity of Critical Load Estimates for Surface Waters to Model Selection and
     Regionalization Schemes, J. Environ. Qual. 22:279-289.
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Holland HD. 1978. The chemistry of the atmosphere and oceans. New York, NY: John Wiley &
       Sons.
Kim, S.W., et al. 2006. Satellite-observed U.S. power plant NOx emission reductions and their 1
       impact on air quality. Geophysical Research Letters 33, L22812, 2
       doi:10.1029/2006GL027749, 2006
Levine, J. S; Bobbe, T; Ray, N; Witt, R. G; Singh, A. 1999.Wildland fires and the environment:
       A global synthesis. (Report no. UNEP/DEIAEW/TR.99.1).
       http://www.na.unep.net/publications/wildfire.pdf. Nairobi, Kenya: Division of
       Environmental Information; Assessment an Early Warning (DEIA&EW); United Nations
       Environment Programme (UNEP).
NARSTO. 2010. Technical Challenges of Multipollutant Air Quality Management; Hidy, G.M.,
       J.R. Brook, K.L. Demerjian, L.T. Molina, W.T. Pennell, and R.D. Scheffe (eds).
       Springer, Dordrecht, 2010.
NARSTO. 2011. Multiple Pollutant Air Quality Management Assessment, Eds. Hidy, G,
       Demerjian, K., Brooke, J, Pennell, W. and R. Scheffe, Springer-Verlag
Norton, S.A.,  J. Kahl, I. Fernandez, T. Haines, L. Rustad, S. Nodvin,.  J. Scofield, T. Stickland,
       H. Erickson, P. Wigington, Jr., and J.Lee. 1999a. The Bear BrookWatershed in Maine
       (BBWM), USA. Environ. Monitor Assess. 55:7-51.
Omernik JM; Powers CF. 1983. Map Supplement: Total Alkalinity of Surface Waters-A
       National Map. Annals of the Association of American Geographers, 73, 133-136
Parton WJ; Hartman M; Ojima D; Schimel D. 1998.  DAYCENT and its land surface submodel:
       Description and testing. Global Planet Change, 19, 35-48.
Parkhurst DL; Appelo CAJ. 1999. User's Guide to PHREEQC (Version 2): A computer program
       for speciation, batch-reaction, one-dimensional transport, and inverse geochemical
       calculations. (Water-Resources Investigations Report). Denver, CO: U.S. Geological
       Survey.
Richter, A., Burrows, J., Nuss, H., Granier,  C. 2005.  Increase to tropospheric nitrogen dioxide 19
       over China observed from space. Nature 437, 129-132..
Rochelle BP; Church MR. 1987. Regional patterns of sulphur retention in watersheds of the
       eastern U.S. Water Air Soil Pollut, 36, 61-73.
Rochelle, B.P. and R. S. Turner. 1992. Direct/Delayed Response Project: Future
       Effects of Long-Term Sulfur Deposition on Stream Chemistry in the Mid-Appalachian
       Region of the Eastern United States. U.S. Environmental Agency. EPA/600/R-92/186.
       Washington, DC. 384 pp.
Russell, A and Samet J.M. 2010. Letter to Administrator Jackson on Review of the Policy
       Assessment of the Secondary National Ambient Air Quality Standard for NOx and SOx :
       First Draft.
Scheffe, R., B. Hubbell, T. Fox, V. Rao and W. Pennell. 2007. The Rationale of a Multiple
       Polluant Multiple Media Air Quality Management Framework, AWMA, EM, May, 2007,
       14-20.
Scheffe, R. , P. Solomon, R.Husar, T. Hanley, J. Rice, M. Schmidt, M. Koerber, M. Gilroy, J.
       Hemby, N. Watkis, M. Papp, J. Rice and J. Tikvart. 2009. The National Ambient Air
       Monitoring Strategy: Rethinking the Role of National Networks, ISSN: 1047-3289 J. Air
       & Waste Manage. Assoc. 59:579-590 DOL10.3155/1047-3289.59.5.579
Sickles, J.E. and D.S. Shadwick. 2007. Changes in air quality and atmospheric deposition in the
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Stoddard J; Kahl JS; Deviney FA; DeWalle DR; Driscoll CT; Herlihy AT; Kellogg JH; Murdoch
       PS; Webb JR; Webster KE. 2003. Response of surface water chemistry to the Clean Air
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       Emission Inventory and Analysis Group; Air Quality and Analysis Division; U.S.
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Warby, R. A. F.,  Johnson, C. E., & Driscoll, C. T. 2009.  Continuing  acidification of organic
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        3      KNOWN OR ANTICIPATED ECOLOGICAL EFFECTS
       In this chapter we address the evaluation of the effects of ambient oxides of nitrogen and
sulfur on ecosystems, and the relationship between those effects and the measure of dose in the
ecosystem, indicated by the depositional loadings of N and S. In section 302(h) of the Clean Air
Act, welfare effects addressed by a secondary NAAQS include, but are not limited to, "effects on
soils, water, crops, vegetation, man-made  materials, animals, wildlife, weather, visibility and
climate, damage to and deterioration of property, and hazards to transportation, as well as effects
on economic values and on personal comfort and well-being". Of these welfare effects
categories, the effects of oxides of nitrogen and sulfur on aquatic and terrestrial ecosystems,
which encompass soils, water, vegetation, wildlife, and contribute to economic value and well-
being, are of most concern at concentrations typically  occurring in the U.S. Direct effects of
oxides of nitrogen and sulfur on vegetation are also discussed in this chapter, and have been the
focus of previous  reviews. However, for this review, the focus of this chapter is on the known
and anticipated effects to ecosystems caused by exposure to oxides of nitrogen and sulfur
through deposition.
       The information presented here is a concise summary of conclusions from the ISA and
the REA. This chapter focuses on effects on specific ecosystems with a brief discussion on
critical uncertainties associated with acidification and  nutrient enrichment.  Those effects are then
evaluated in Chapter 4 within the context of alternative definitions of, including assessments of
potential impacts  on ecosystem services.  Effects are broadly categorized into acidification and
nutrient-enrichment in the proceeding sections.  This is background information intended to
support new approaches for the design of ecologically relevant secondary oxides of nitrogen and
sulfur standards which are protective of U.S. ecosystems. More detailed information on the
conceptual design and specific options for the standards  is presented in Chapters 6  and 7 of this
policy assessment document. While we provide a summary of effects for four of the primary
effects categories  (aquatic acidification, terrestrial acidification, aquatic nutrient enrichment, and
terrestrial nutrient enrichment), we reiterate that the focus of this second draft policy assessment
is on effects related to aquatic acidification, without downplaying the potential importance of
effects in other categories.
                                           3-1

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3.1    ACIDIFICATION: EVIDENCE OF EFFECTS ON STRUCTURE AND
       FUNCTION OF TERRESTRIAL AND FRESHWATER ECOSYSTEMS
   Sulfur oxides (SOX) and nitrogen oxides (NOx) in the atmosphere undergo a complex mix of
reactions in gaseous, liquid, and solid phases to form various acidic compounds. These acidic
compounds are removed from the atmosphere through deposition: either wet (e.g., rain, snow),
fog or cloud, or dry (e.g., gases, particles). Deposition of these acidic compounds to ecosystems
can lead to effects on ecosystem structure and function. Following deposition, these compounds
can, in some instances unless retained by soil or biota, leach out of the soils in the form of sulfate
(SO42) and nitrate (NO3-), leading to the acidification of surface waters. The effects on
ecosystems depend on the magnitude and rate of deposition, as well as a host of biogeochemical
processes occurring in the soils and waterbodies (REA 2.1). The chemical forms of nitrogen that
may contribute to acidifying deposition include both oxidized and reduced chemical species.
       When sulfur or nitrogen leaches from soils to surface waters in the form of SO42 or NO3,
an equivalent amount of positive cations, or countercharge,  is also transported. This maintains
electroneutrality. If the countercharge is provided by base cations, such as calcium (Ca2+),
magnesium (Mg2+), sodium (Na+), or potassium (K+), rather than hydrogen (H+) and dissolved
inorganic  aluminum, the acidity of the soil water is neutralized, but the base saturation of the soil
decreases. Continued SO42 or NO3-leaching can deplete the available base cation pool in soil. As
the base cations are removed, continued deposition and leaching of SO42 and/or NO3-(with H+and
A13+) leads to acidification of soil water, and by connection,  surface water. Introduction of strong
acid anions such as sulfate and nitrate to an already  acidic soil, whether naturally or due to
anthropogenic activities, can lead to instantaneous acidification of waterbodies through direct
runoff without any significant change in base cation saturation. The ability of a watershed to
neutralize acidic deposition is determined by a variety of biogeophysical factors including
weathering rates, bedrock composition, vegetation and microbial processes, physical and
chemical characteristics of soils and hydrologic flowpaths. (REA 2.1)  Some of these factors
such as vegetation and soil depth are highly  variable over small spatial scales such as meters, but
can be aggregated to evaluate patterns over larger spatial scales. Acidifying deposition of oxides
of nitrogen and sulfur and the chemical and biological responses associated with these inputs
vary temporally. Chronic or long-term deposition processes in the time scale of years to decades
result in increases in inputs of N and S to ecosystems and the associated ecological effects.
                                           3-2

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 Episodic or short term (i.e., hours or days) deposition refers to events in which the level of the
 acid neutralizing capacity (ANC) of a lake or stream is temporarily lowered.  In aquatic
 ecosystems, short-term (i.e., hours or days) episodic changes in water chemistry can have
 significant biological effects.  Episodic acidification refers to conditions during precipitation or
 snowmelt events when proportionately more drainage water is routed through upper soil horizons
 that tend to provide less acid neutralizing than was passing through deeper soil horizons (REA
 4.2). In addition, the accumulated sulfate and nitrate in snow packs can provide a surge of acidic
 inputs. Some streams and lakes may have chronic or base flow chemistry that is suitable for
 aquatic biota, but may be subject to occasional acidic episodes with deleterious consequences to
 sensitive biota.
        The following summary is a concise overview of the known or anticipated effects caused
 by acidification to ecosystems within the United States.  Acidification affects both terrestrial and
 freshwater aquatic ecosystems.

3.1.1  Nature of acidification-related ecosystem responses
        The ISA concluded that deposition of SOX, NOX, and NHX leads to the varying degrees of
 acidification of ecosystems (EPA 2008). In the process of acidification, biogeochemical
 components of terrestrial and  freshwater aquatic ecosystems are altered in a way that leads to
 effects on biological organisms. Deposition to terrestrial ecosystems often moves through the
 soil and eventually leaches into adjacent water bodies.
Aquatic ecosystems
        The scientific evidence is sufficient to infer a causal relationship between acidifying
 deposition and effects on biogeochemistry and biota in aquatic ecosystems (ISA 4.2.2). The
 strongest evidence comes from studies of surface water chemistry in which acidic deposition is
 observed to alter sulfate and nitrate concentrations in surface waters, the sum of base cations,
 ANC, dissolved inorganic aluminum and pH. (ISA 3.2.3.2).  Consistent and coherent
 documentation from multiple  studies on various species from all  major trophic levels of aquatic
 systems shows that geochemical alteration caused by acidification can result  in the loss of acid-
 sensitive biological species (ISA 3.2.3.3).  For example,  in the Adirondacks, of the 53 fish
 species recorded in Adirondack lakes about half (26 species) were absent from lakes with pH
 below 6.0 (Baker et al, 1990). Biological effects are linked to changes in water chemistry
                                            3-3

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including decreases in ANC and pH and increases in inorganic Al concentration. The direct

biological effects are caused by lowered pH which leads to in increased inorganic Al

concentrations (Figure 3-1 and 3-2). While ANC level does not cause direct biological harm it is

a good overall indicator of the risk of acidification (See further discussion in Section 3.1.3).
                     t NO3-andSO42-
                                                        Stream water chemistry
                                      I
        ^Ecological effects which may include:

        Lower biodiversity
        Altered species composition
        Localized extinction of sensitive species
        Individual mortality of sensitive species
        Sub-lethal effects to sensitive species and
        ecological integrity
Aquatic biota
Figure 3-1.  Conceptual model of direct and indirect acidification effects on aquatic biota.
Acidic pollutants (NO3- and SO4-2) lower ANC, resulting in lower pH with direct toxic effects
on fish. The lower pH mobilizes A13+ from soils often resulting in higher concentration in
stream water causing direct toxicity to fish.
                                          3-4

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        o
           10
                                              pH
                                                      8
10
12
Figure 3-2.  Equilibrium diagram of the log of aluminum species concentrations as a function of
pH.  At lower pH there is far greater proportion of dissolved Al species relative to solid gibbsite
(solid aluminum hydroxide, A1(OH)3) .   The most toxics form of Al, free trivalent aluminum,
(Al+3) levels rapidly increase with lower pH values.
       There are clear associations between pH and aquatic species mortality and health which
are summarized in Tables 3-1 and 3-2 and illustrated in Figure 3-3.  Significant harm to sensitive
aquatic species has been observed at pH levels below 6.  Normal stream pH levels with little to
no toxicity ranges from 6 to 7 (MacAvoy et al, 1995).  Baker et al (1990) observed that "lakes
with pH less than approximately 6.0 contain significantly fewer species than lakes with pH levels
above 6.0".  As noted in Chapter 3, typically at pH <4.5 and an ANC <0 ueq/L, complete to
near-complete loss of many taxa of organisms occur, including fish and aquatic insect
populations, whereas other taxa are reduced to only acidophilic species.
       Additional evidence can help refine the understanding of effects occurring at pH levels
between 4.5 and 6. When pH levels are below 5.6, relatively lower trout survival rates were
observed in the Shenendoah National Park. In field observations, when pH levels dropped to 5,
mortality rates went to 100 percent. (Bulger et al, 2000). At pH levels ranging from 5.4 to 5.8,
                                           3-5

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cumulative mortality continues to increase. Several studies have shown that trout exposed to
water with varying pH levels and fish larvae showed increasing mortality as pH levels decrease.
In one study almost 100 percent mortality was observed at a pH of 4.5 compared to almost 100
percent survival at a pH of 6.5. Intermediate pH values (6.0, 5.5) in all cases showed reduced
survival compared with the control (6.5), but not by statistically significant amounts (ISA
3.2.3.3).
       One important indicator of acid stress is increased fish mortality.  The response offish to
pH is not uniform across species. A number of synoptic surveys indicated loss of species
diversity and absence of several fish species in the pH range of 5.0 to 5.5.  If pH is lower, there is
a greater likelihood that more fish species could be lost without replacement, resulting in
decreased richness and diversity. In general, populations of salmonids are not found at pH levels
less than 5.0, and smallmouth bass (Micropterus dolomieu) populations are usually not found at
pH values less than about 5.2 to 5.5. From Table 3-1,  only one study showed significant
mortality effects above  a pH of 6, while a number of studies showed significant mortality when
pH levels are at or below 5.5.
       The highest pH level for any of the studies reported in Table 3-2 is 6.0, suggesting that
pH above 6.0 is protective against mortality effects for most species.  Most thresholds are in the
range of pH of 5.0 to 6.0, which suggests that a target  pH should be no lower than 5.0.
Protection against mortality in some recreationally important species such as lake trout (pH
threshold of 5.6) and crappie (pH threshold of 5.5), combined with the evidence of effects on
larval and embryo  survival suggests that pH levels greater than 5.5 should be targeted to provide
protection against mortality effects throughout the life stages offish.
      Non-lethal effects have been observed  at pH levels as high as 6.  A study in the
Shenendoah National Park found that the condition factor, a measure offish health expressed as
fish weight/length3 multiplied by a scaling constant, is positively correlated with stream pH
levels, and that the condition factor is reduced in streams with a pH of 6.0 (ISA 3.2.3.3).
      Biodiversity is another indicator of aquatic ecosystem health.  As discussed in Chapter 2,
a key study in the Adirondacks found that lakes with a pH of 6.0 had only half the potential
species offish (27  of 53 potential species).  There is often a positive relationship between pH and
number offish species,  at least for pH values between about 5.0 and 6.5, or ANC values between
about 0 to 100 jieq/L (Bulger et al, 1999; Sullivan et al, 2006). Such observed relationships are
                                           3-6

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complicated, however, by the tendency for smaller lakes and streams, having smaller watersheds,
to also support fewer fish species, irrespective of acid-base chemistry. This pattern may be due to
a decrease in the number of available niches as stream or lake size decreases. Nevertheless, fish
species richness is relatively easily determined and is one of the most useful indicators of
biological effects of surface water acidification.
       In a study of Ontario lakes, Matuszek and Beggs (1988) found that the number offish
species is positively correlated with pH, with a clear loss of species starting at pH levels less than
or equal to 5.5. This relationship is displayed in Figure 3-3.
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Table 3-1. Summary of Fish Mortality Response to pH; source: EPA, 2008 (ISA).
Mortality
Endpoint
Increased
Mortality
>50% larval
mortality
embryo
survival
Significant
decrease
>50%
embryo
mortality
Substantial
reduction
Authors
Johnson et al.
(1987)
Holtze and
Hutchinson
(1989)
Johansson et al.
(1977)
Swenson et al.
(1989)
Mills et al. (1987)
Buckler et al.
(1987)
Klauda et al.
(1987)
Kane and Rabeni
(1987)

McCormick et al.
(1989)
Holtze and
Hutchinson
(1989)
Baker and
Schofield (1980)
Species
Blacknose dace,
creek chub
Brook trout
Common shiner
Lake whitefish,
white sucker,
walleye
Smallmouth bass
Atlantic salmon
Brown trout
Brook Trout
Black crappie
Rock bass
Yellow perch,
largemouth bass
Fathead minnow
Slimy sculpin
Lake Trout
Pearl dace
White sucker
Striped bass
Blueback herring
Smallmouth bass

Fathead minnow
Common shiner
White sucker
pH
Level
5.9-6.0
4.8-5.1
5.4-6.0
5.1 -5.2
4.8
5.0
4.5-5.0
4.5
5.5
5.0
4.5
5.9
5.6-5.9
5.6
5.1
5.0-5.1
6.5
5.7
5.1

6.0
5.4
5.2
Notes
In situ bioassay with early
life stages in Adirondack
surface waters
Laboratory exposure of
early life stages to pH and
Al.
Laboratory tests with eggs
exposed to low pH, no Al.
Laboratory tests with early
life stages exposed to pH
and Al.
Whole-lake treatment (fish
population recruitment
failure)
Lab bioassay
Lab bioassay
Lab bioassay

Lab bioassay
Lab bioassay
Lab bioassay
                                       3-8

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      Table 3-2. General summary of biological changes anticipated with surface water
	acidification, expressed as a decrease in surface water pH.	
    pH                               General Biological Effects
 Decrease
 6.5 to 6.0   Small decrease in species richness of plankton and benthic invertebrate
            communities resulting from the loss of a few highly acid-sensitive species, but no
            measurable change in total community abundance or production.
            Some adverse  effects (decreased reproductive success) may occur for highly
            acid-sensitive  fish species (e.g.,  fathead minnow, striped bass).
 6.0 to 5.5   Loss of sensitive species of minnows and dace, such as fathead minnow and
            blacknose dace; in some waters, decreased reproductive success of lake trout and
            walleye, which are important sport fish species in some areas.
            Visual accumulation of filamentous green algae in the near-shore zone of many
            lakes and in some streams.
            Distinct decrease in species richness and change in species composition of
            plankton and benthic invertebrate communities, although little if any change in
            total community abundance or production.
            Loss of some common invertebrate species from zooplankton and benthic
            communities, including many species of snails, clams, mayflies, and amphipods,
            and some crayfish.
 5.5 to 5.0   Loss of several important sport fish species, including lake trout, walleye,
            rainbow trout, and smallmouth bass, as well as additional nongame species such
            as creek chub.
            Further increase in the extent and abundance of filamentous green algae in lake
            near-shore areas and streams.
            Continued shift in species composition and decline in species richness of
            plankton, periphyton, and benthic invertebrate communities; decreases in total
            abundance and biomass  of benthic invertebrates and zooplankton may occur in
            some waters.
            Loss of several additional invertebrate species common in surface waters,
            including all snails,  most species of clams, and many species of mayflies,
            stoneflies, and other benthic invertebrates.
            Inhibition of nitrification.
 5.0 to 4.5   Loss of most fish species, including most important sport fish species such as
            brook trout and Atlantic salmon. A few fish species are able to survive and
            reproduce in water below pH 4.5 (e.g., central mudminnow, yellow perch, and in
            some waters, largemouth bass).
            Measurable decline  in the whole-system rates of decomposition of some forms of
            organic matter, potentially resulting in decreased rates of nutrient cycling.
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    +2
 (O
 LU
 8>
 OL
 LU
 OQ
 LLI
 o:
     -1
     -2
     -3
     -4
-5
 2-6
     -7
            4.5
                5.0
5.5
6.0
6.5
PH
7.0
7.5
8.0
8.5
Figure 3-3. Mean residual number of species per lake for lakes in Ontario, by pH interval. The
residual number of species for a lake is the deviation of the observed number from the number
predicted by lake area (Matuszek and Beggs,1988).

       Changes in stream water pH also contribute to declines in taxonomic richness of
zooplankton, and  macroinvertebrates which are often sources of food for fish, birds and other
animal species in various ecosystems. These fish may also serve as a source of food and
recreation for humans (see Chapter 4). Acidification of ecosystems has been shown to disrupt
food web dynamics causing alteration to the diet, breeding distribution and reproduction of
certain species of birds (ISA Section 4.2.2.2. and Table 3-9).  For  example, breeding distribution
of the common goldeneye (Bucephala clanguld) an insectivorous  duck may be affected by
changes in acidifying deposition (Longcore and Gill, 1993). Similarly, decreases in prey
diversity and quantity have been observed to create feeding problems for nesting pairs of loons
on low-pH lakes in the Adirondacks (Parker 1988).
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Terrestrial ecosystems
       In terrestrial ecosystems, the evidence is sufficient to infer a causal relationship between
acidifying deposition and changes in biogeochemistry (ISA 4.2.1.1). The strongest evidence
comes from studies of forested ecosystems, with supportive information on other plant taxa,
including shrubs and lichens (ISA 3.2.2.1.).  Three useful indicators of chemical changes and
acidification effects on terrestrial ecosystems, showing consistency and coherence  among
multiple studies are: soil base saturation, Al concentrations in soil water and soil C:N ratio (ISA
3.2.2.2).
       In soils with base saturation less than about 15 to 20%, exchange  chemistry is dominated
by Al (Reuss,  1983).  Under these conditions, responses to inputs of sulfuric acid and nitric acid
largely involve the release and mobilization of dissolved inorganic Al.  The effect can be
neutralized by weathering from geologic parent material or base cation exchange. The Ca2+ and
Al concentrations in soil water are strongly influenced by soil acidification and both have been
shown to have quantitative links to tree health, including Al interference  with Ca2+ uptake and Al
toxicity to roots (Parker et al, 1989; U.S. EPA, 2009). Effects of nitrification and  associated
acidification and cation leaching have been consistently shown to occur only in soils with a C:N
ratio below about 20 to 25 (Aber et al., 2003; Ross et al., 2004).
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       Soil acidification caused by acidic deposition has been shown to cause decreased growth
      and increased susceptibility to disease and injury in sensitive tree species. Red spruce
      (Picea rubens) dieback or decline has been observed across high elevation areas in the
      Adirondack, Green and White mountains (DeHayes et al,  1999).  The frequency of
      freezing injury to red spruce needles has increased over the past 40 years, a period that
      coincided with increased emissions of S and N oxides and  increased acidifying deposition
      (DeHayes et al.,  1999).  Acidifying deposition can contribute to dieback in sugar maple
      (Acer saccharum) through depletion of cations from soil with low levels of available Ca
      (Horsley et al., 2000; Bailey et al., 2004). Grasslands are likely less sensitive to
      acidification than forests due to grassland soils being generally rich in base cations (Fenn
      et al., 2003; Blake  et al., 1999).

3.1.2 Ecosystem sensitivity to acidification
       The intersection between current deposition loading, historic loading,  and sensitivity
defines the ecological vulnerability to the effects of acidification. Freshwater  aquatic and some
terrestrial ecosystems,  notably forests, are the ecosystem types which are most sensitive to
acidification.  The ISA reports that the principal factor governing the sensitivity of terrestrial and
aquatic ecosystems to acidification from sulfur and nitrogen deposition is geology (particularly
surficial geology). Geologic formations having low base cation supply generally underlie the
watersheds of acid-sensitive lakes  and streams. Other factors that contribute to the  sensitivity of
soils and surface waters to acidifying deposition include topography, soil chemistry, land use,
and hydrologic flowpaths. Episodic and chronic acidification tends to occur in areas that have
base-poor bedrock, high  relief, and shallow soils (ISA 3.2.4.1).

3.1.3 Magnitude of ecosystem responses to acidifying deposition
       Terrestrial and  aquatic ecosystems differ in their response to acidifying deposition.
Therefore the magnitude of ecosystem response is described separately for aquatic  and terrestrial
ecosystems in the following sections.  The magnitude of response refers to both the severity of
effects and the spatial extent of the U.S. which is affected.
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Aquatic acidification
       Freshwater ecosystem surveys and monitoring in the eastern United States have been
conducted by many programs since the mid-1980s, including EPA's Environmental Monitoring
and Assessment Program (EMAP), National Surface Water Survey (NSWS), Temporally
Integrated Monitoring of Ecosystems (TIME) (Stoddard, 1990), and Long-term Monitoring
(LTM) (Ford et al, 1993; Stoddard et al, 1996) programs. Based on analyses of surface water
data from these programs, New England, the Adirondack Mountains, the Appalachian Mountains
(northern Appalachian Plateau and Ridge/Blue Ridge region), and the Upper Midwest contain
the most sensitive lakes and streams (i.e., ANC less than about 50 ueq/L). Portions of northern
Florida also contain many acidic and low-ANC lakes and streams, although the role of acidifying
deposition in this region is less clear. The western U.S.  contains many of the surface waters most
sensitive to potential acidification  effects, but with the exception of the Los Angeles Basin and
surrounding areas, the levels of acidifying deposition are low in most  areas.  Therefore,
acidification of surface waters by acidic deposition is not as prevalent in the western U.S., and
the extent of chronic surface water acidification that has occurred in that region to date has likely
been very limited relative to the Eastern U.S. (ISA 3.2.4.2  and REA 4.2.2).
       There are a number of species including fish, aquatic insects, other invertebrates and
algae that are sensitive to acidification and cannot survive, compete, or reproduce in acidic
waters (ISA 3.2.3.3). Decreases in ANC and pH have been shown to contribute to declines in
species richness and declines in abundance of zooplankton, macroinvertebrates, and fish (Keller
and Gunn 1995; Schindler et al., 1985). Reduced growth rates have been attributed to acid stress
in a number offish species including Atlantic salmon (Salmo salar), Chinook salmon
(Oncorhynchus tshawytscha), lake trout (Salvelinus namaycush), rainbow trout (Oncorhynchis
mykiss), brook trout (Salvelinus Fontinalis\ and brown trout (Salmo trutta) (Baker et al.,  1990).
In response to small to moderate changes in acidity, acid-sensitive species are often replaced by
other more acid-tolerant species, resulting in changes in community composition and richness.
The effects of acidification are continuous, with more species being affected at higher degrees of
acidification.  At a point, typically a pH <4.5 and an ANC <0 [j,eq/L,  complete to near-complete
loss of many taxa of organisms occur, including fish and aquatic insect populations, whereas
other taxa are reduced to only acidophilic species. These changes in taxa composition are
associated with the high energy cost in maintaining physiological homeostasis, growth, and
                                          3-13

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reproduction at low ANC levels (Schreck, 1981,  1982; Wedemeger et al, 1990; REA appendix
2.3). Decreases in species richness related to acidification have been observed in the Adirondack
Mountains and Catskill Mountains of New York (Baker et al., 1996), New England and
Pennsylvania (Haines and Baker, 1986), and Virginia (Bulger et al., 2000). From the sensitive
areas identified by the ISA, further "case study" analyses on aquatic ecosystems in the
Adirondack Mountains  and Shenandoah National Park were conducted to better characterize
ecological risk associated with acidification (REA Chapter 4).
       ANC is the most widely used indicator of acid sensitivity and has been found in various
studies to be the best single indicator of the biological response and health of aquatic
communities in acid-sensitive systems (Lien et al., 1992; Sullivan et al., 2006; EPA, 2008).  In
the REA, surface water trends in SC>42" and N(V concentrations and ANC levels were analyzed
to affirm the understanding that reductions in deposition could influence the risk of acidification.
ANC values were categorized according to their effects on biota, as shown in Table 3-3.
Monitoring data from TIME/LTM and EMAP programs were assessed for the years 1990 to
2006, and past, present, and future water quality levels were estimated by both steady-state and
dynamic biogeochemical models.
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        Table 3-3.  Ecological effects associated with alternative levels of acid
        neutralizing capacity (ANC). (source: USEPA, Acid Rain Program)
                   Category Label ANC Levels and Expected Ecological Effects
  Acute
  Concern
<0 ueq/L
Complete loss of fish populations is expected. Planktonic communities
have extremely low diversity and are dominated by acidophilic taxa.
The numbers of individuals in plankton species that are present are
greatly reduced.
  Severe
  Concern
0-20 ueq/L
Highly sensitive to episodic acidification. During episodes of high
acidifying deposition, brook trout populations may experience lethal
effects. The diversity and distribution of zooplankton communities
decline sharply.
  Elevated
  Concern
20-50 ueq/L
Fish species richness is greatly reduced (i.e., more than half of expected
species can be missing). On average, brook trout populations
experience sublethal effects, including loss of health, ability to
reproduce, and fitness. Diversity and distribution of zooplankton
communities decline.
  Moderate
  Concern
50-100
ueq/L
Fish species richness begins to decline (i.e., sensitive species are lost
from lakes). Brook trout populations are sensitive and variable, with
possible sublethal effects. Diversity and distribution of zooplankton
communities also begin to decline as species that are sensitive to
acidifying deposition are affected.
  Low
  Concern
>100 ueq/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.
        Studies on fish species richness in the Adirondacks Case Study Area demonstrated the

effect of acidification. Of the 53 fish species recorded in Adirondack Case Study Area lakes,

only 27 species were found in lakes with a pH <6.0. The 26 species missing from lakes with a

pH <6.0 include important recreational species, such as Atlantic salmon, tiger trout (Salmo trutta
X Salvelinus fontinalis), redbreast sunfish (Lepomis auritus), bluegill (Lepomis macrochims),

tiger musky (Esox masquinongy X lucius), walleye (Sander vitreus), alewife (Alosa

pseudoharengus), and kokanee (Oncorhynchus nerkd) (Kretser et al, 1989), as well as

ecologically important minnows that are commonly consumed by sport fish. A survey of 1,469

lakes in the late 1980s found 346 lakes to be devoid offish. Among lakes with fish, there was a
relationship between the number offish species and lake pH, ranging from about one species per

lake for lakes having a pH <4.5 to about six species per lake for lakes having a pH >6.5 (Driscoll
                                             3-15

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et al, 2001; Kretser et al, 1989). In the Adirondacks, a positive relationship exists between the
pH and ANC in lakes and the number offish species present in those lakes (ISA 3.2.3.4).
       Since the mid-1990s, streams in the Shenandoah Case Study Area have shown slight
declines in NOs- and SC>4 2" concentrations in surface waters. The 2006 concentrations are still
above pre-acidification (1860) conditions. MAGIC modeling predicts surface water
concentrations of MV and SC>42" arelO- and 32-fold higher, respectively, in 2006 than in 1860.
The estimated average ANC across 60 streams in the Shenandoah Case Study Area is 57.9 ueq/L
(±4.5 ueq/L). 55% of all monitored streams in the Shenandoah Case Study Area have a current
risk of Elevated, Severe, or Acute. Of the 55%, 18% are chronically acidic today (REA 4.2.4.3).
       Based on a deposition scenario for this study area that maintains current emission levels
from 2020 to 2050, the simulation forecast indicates that a large number of streams still have
Elevated to Acute problems with acidity. In fact, from 2006 to 2050, the percentage of streams
with Acute Concern is predicted to increase by 5%, while the percentage of streams in Moderate
Concern decreases by 5%.
       Biological effects of increased acidification documented in the Shenandoah Case Study
Area include a decrease in the condition factor in blacknose dace (Dennis and Bulgar 1999,
Bulgar et al., 1999) and a decrease in fish biodiversity associated with decreasing stream ANC
(Bulger et al.,1995; Dennis and Bulger, 1999; Dennis et al., 1995; MacAvoy and Bulger, 1995,
Bulgar et al., 1999).  On average, the fish species richness is lower by one fish  species for every
21 ueq/L decrease in ANC in Shenandoah National Park streams (ISA 3.2.3.4).

Terrestrial acidification
       The ISA identified a variety of indicators that can be used to measure the effects of
acidification in soils. Most effects of terrestrial acidification are observed in sensitive forest
ecosystem in the U.S. Tree health has been linked to the availability of base cations (Be) in soil
(such as Ca2+, Mg2+ and potassium), as well as soil Al content. Tree species show a range of
sensitivities to Ca/Al and Bc/Al soil molar ratios, therefore these are good chemical indicators
because they directly relate to the biological effects. Critical Bc/Al molar ratios for a large
variety of tree species ranged from 0.2 to 0.8 (Sverdrup and Warfvinge, 1993, a meta-data
analysis of laboratory and field studies). This range is similar to critical ratios of Ca/Al. Plant
                                           3-16

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toxicity or nutrient antagonism was reported to occur at Ca/Al molar ratios ranging from 0.2 to
2.5 (Cronan and Grigal, 1995; meta-data assessment) (REA pg 4-54, REA Appendix 5).
       There has been no systematic national survey of terrestrial ecosystems to determine the
extent and distribution of terrestrial ecosystem sensitivity to the effects of acidifying deposition.
However, one preliminary national evaluation estimated that -15% of forest ecosystems in the
U.S. exceed the estimated critical load based on soil ANC leaching for S and N deposition by
>250 eq ha"1 yr"1 (McNulty et al, 2007). Forests of the Adirondack Mountains of New York,
Green Mountains of Vermont, White Mountains of New Hampshire, the Allegheny Plateau of
Pennsylvania, and high-elevation forest ecosystems in the southern Appalachians are the regions
most sensitive to terrestrial acidification effects from acidifying deposition (ISA 3.2.4.2). While
studies show some recovery of surface waters, there are widespread measurements of ongoing
depletion of exchangeable base cations in forest soils in the northeastern U.S. despite recent
decreases in acidifying deposition, indicating a slow recovery time.
       In the REA, a critical load analysis was performed for sugar maple and red spruce forests
in the eastern United States by using Bc/Al ratio in acidified forest soils as an indicator to assess
the impact of nitrogen and sulfur deposition on tree health. These are the two most commonly
studied tree  species in North America for effects of acidification. At  a Bc/Al ratio of 1.2, red
spruce growth can be decreased by 20%. Sugar maple growth can be decreased by 20% at  a
Bc/Al ratio of 0.6 (REA 4.4). The REA analysis determined the health of at least a portion of the
sugar maple and red spruce growing in the United States may have been compromised with
acidifying total nitrogen and sulfur deposition. Specifically, total nitrogen and sulfur deposition
levels exceeded three selected critical loads for tree growth in 3% to  75% of all sugar maple
plots across  24 states. For red spruce, total nitrogen and sulfur deposition levels exceeded three
selected critical loads in 3% to 36% of all red spruce plots across eight states (REA 4.4).

3.1.4 Key uncertainties associated with acidification
       There are different levels of uncertainty associated with relationships between deposition,
ecological effects and ecological indicators. In Chapter 7 of the REA, the case study analyses
associated with each targeted effect area were synthesized by identifying the strengths,
limitations, and uncertainties associated with the available data, modeling approach, and
relationship  between the selected ecological indicator and atmospheric deposition as described
by the ecological effect function (REA Figure  1-1). The key uncertainties were characterized as
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follows to evaluate the strength of the scientific basis for setting a national standard to protect
against a given effect (REA 7.0):
    •   Data Availability: high, medium or low quality. This criterion is based on the
       availability and robustness of data sets, monitoring networks, availability of data that
       allows for extrapolation to larger assessment areas, and input parameters for modeling
       and developing the ecological effect function. The scientific basis for the ecological
       indicator selected is also incorporated into this criterion.
    •   Modeling Approach: high, fairly high, intermediate, or low confidence.  This value is
       based on the strengths and limitations of the models used in the analysis and how
       accepted they are by the scientific community for their application in this analysis.
    •   Ecological Effect Function: high, fairly high, intermediate, or low confidence. This
       ranking is based  on how well the ecological effect function describes the relationship
       between atmospheric deposition and the ecological indicator of an effect.
Aquatic acidification
       The REA concludes that the available data are robust and considered high quality. There
is high confidence about the use of these data and their value for extrapolating to a larger
regional population of lakes. The EPA TIME/LTM network represents a source of long-term,
representative sampling.  Data on sulfate concentrations, nitrate concentrations and ANC from
1990 to 2006 used for this analysis as well as EPA EMAP and REMAP surveys, provide
considerable data on surface water trends.
       There is fairly high confidence associated with modeling and input parameters.
Uncertainties in water quality estimates (i.e., ANC) from MAGIC were derived from multiple
site calibrations. The  95% confidence interval for pre-acidification of lakes was an average of 15
|ieq/L difference in ANC concentrations or 10% and 8 |ieq/L or 5% for streams (REA 7.1.2).
The use of the critical load model used to estimate aquatic critical loads is limited by the
uncertainties associated with runoff and surface water measurements and in estimating the
catchment supply of base cations from the weathering of bedrock and soils (McNulty et al,
2007).  To propagate uncertainty in the model parameters, Monte Carlo methods were employed
to develop an inverse function of exceedences. There is high confidence associated with the
ecological effect function developed for aquatic acidification. In calculating the ANC function,
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the depositional load for N or S is fixed by the deposition of the other, so deposition for either
will never be zero (REA Figure 7.1-6).
Terrestrial acidification
       The available data used to quantify the targeted effect of terrestrial acidification are
robust and considered high quality. The USFS-Kane Experimental Forest and significant
amounts of research work in the Allegheny Plateau have produced extensive, peer-reviewed data
sets. A meta-analysis of laboratory studies showed that tree growth was decreased by 20%
relative to controls for BC/A1  ratios (ISA 7.2.1 and Figure 7.2-1).  Sugar maple and red spruce
were the focus of the REA since they are demonstrated to be negatively affected by soil available
Ca2+ depletion and high concentrations of available Al, and occur in areas that receive high
acidifying deposition, There is high confidence about the use of the REA terrestrial acidification
data and their value for extrapolating to a larger regional population of forests.
       There is high confidence associated with the models, input parameters, and assessment of
uncertainty used in the case study for terrestrial acidification. The Simple Mass Balance (8MB)
model, a commonly used and  widely applied approach for estimating critical loads, was used in
the REA analysis (ISA 7.2.2).  There is fairly high confidence associated with the ecological
effect function developed for terrestrial acidification (REA 7.2.3).

3.2    NITROGEN ENRICHMENT: EVIDENCE OF EFFECTS ON STRUCTURE AND
       FUNCTION OF TERRESTRIAL AND FRESHWATER ECOSYSTEMS
       The following summary is a concise overview of the known or anticipated effects caused
by nitrogen nutrient enrichment to ecosystems within the United States. Nutrient-enrichment
affects terrestrial, freshwater and estuarine ecosystems. Nitrogen deposition is a major source of
anthropogenic nitrogen.  For many terrestrial and freshwater ecosystems other sources of
nitrogen including fertilizer and waste treatment  are greater than deposition.  Nitrogen deposition
often contributes to nitrogen-enrichment effects in estuaries, but does not drive the effects since
other sources of N greatly exceed N deposition.  Both oxides of nitrogen and reduced forms of
nitrogen (NHx) contribute to nitrogen deposition. For the most part, nitrogen effects on
ecosystems do not depend on  whether the nitrogen is in oxidized or reduced form.  Thus, this
summary focuses on the effects of nitrogen deposition in total.
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3.2.1   Nature of nutrient enrichment-related ecosystem responses
       The ISA found that deposition of nitrogen, including NOX and NHX, leads to the nitrogen
enrichment of ecosystems (EPA 2008).  In the process of nitrogen enrichment, biogeochemical
components of terrestrial and freshwater aquatic ecosystems are altered in a way that leads to
effects on biological organisms.
       The evidence is sufficient to infer a causal relationship between N deposition and the
alteration of biogeochemical cycling in terrestrial ecosystems (ISA 4.3.1.1 and 3.3.2.1). This is
supported by numerous observational, deposition gradient and field addition experiments in
sensitive ecosystems. Stoddard (1994) identified the leaching of NOs- in soil drainage waters and
the export of NOs- in steam water as two of the primary indictors of N enrichment.  Several N-
addition studies indicate that NOs- leaching is induced by chronic additions of N (Edwards et al,
2002; Kahl et al.,  1999; Peterjohn et al., 1996; Norton et al., 1999). Aber et al. (2003) found that
surface water N(V concentrations exceeded 1 |ieq/L in  watersheds receiving about 9 to 13 kg
N/ha/yr of atmospheric N deposition. N deposition disrupts the nutrient balance of ecosystems
with numerous biogeochemical effects.  The chemical indicators that are typically measured
include NOs  leaching, soil C:N ratio, rates of N mineralization, nitrification, denitrification,
foliar N concentration, and soil water NOs - and NH4+ concentrations. Note that N saturation (N
leaching from ecosystems) does not need to occur to cause effects. Substantial leaching of NOs-
from forest soils to stream water can acidify downstream waters, leading to effects described in
the previous section on aquatic acidification. Due to the complexity of interactions between the
N and C cycling, the effects of N on C budgets (quantified input and output of C to the
ecosystem) are variable.  Regional trends in net ecosystem productivity (NEP) of forests (not
managed for silviculture) have been estimated through models based on gradient studies and
meta-analysis. Atmospheric N deposition has been shown to cause increased litter accumulation
and carbon storage in above-ground woody biomass (Thomas et. al., 2010).  In the West, this has
lead to increased susceptibility to more  severe fires. Less is known regarding the effects of N
deposition on C budgets of non-forest ecosystems.
       The evidence is sufficient to infer a causal relationship between N deposition on the
alteration of species richness, species composition and biodiversity in terrestrial ecosystems
(ISA 4.3.1.2).  Some organisms and ecosystems are more sensitive to N deposition and effects of
N deposition are not observed in all habitats.  The most  sensitive terrestrial taxa to N deposition
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are lichens. Empirical evidence indicates that lichens in the U.S. are affected by deposition levels
as low as 3 kg N/ha/yr. Alpine ecosystems are also sensitive to N deposition, changes in an
individual species (Carex rupestris) were estimated to occur at deposition levels near 4 kg N
/ha/yr and modeling indicates that deposition levels near 10 kg N/ha/yr alter plant community
assemblages. In several grassland ecosystems, reduced species diversity and an increase in non-
native, invasive species are associated with N deposition (Clark and Tillman, 2008; Schwinning
et al, 2005).
       In freshwater ecosystems, the evidence is sufficient to infer a causal relationship between
N deposition and the alteration of biogeochemical cycling in freshwater aquatic ecosystems (ISA
3.3.2.3). N deposition is the main source of N enrichment to headwater streams, lower order
streams and high  elevation lakes. The most common chemical indicators that were studied
included NOs  and dissolved inorganic nitrogen (DIN) concentration in surface waters as well as
Chi a:total P ratio. Elevated surface water NOs  concentrations occur in both the eastern and
western U.S. Bergstrom and Jansson (2006) report a significant correlation between N deposition
and lake biogeochemistry by identifying a correlation between wet deposition and [DIN] and Chi
a: Total P. Recent evidence provides examples of lakes  and streams that are limited by N and
show signs of eutrophication in response to N addition.
       The evidence is sufficient to infer a causal relationship between N deposition and the
alteration of species richness, species composition and biodiversity in freshwater aquatic
ecosystems (ISA 3.3.5.3). Increased N deposition can cause a shift in community composition
and reduce algal biodiversity, especially in sensitive oligotrophic lakes.
       In the ISA, the evidence is sufficient to infer a causal relationship between Nr deposition
and the biogeochemical cycling of N and carbon (C) in  estuaries (ISA 4.3.4.1 and 3.3.2.3). In
general, estuaries tend to be nitrogen-limited, and many currently receive high levels of nitrogen
input from human activities (REA 5.1.1). It is unknown if atmospheric deposition alone is
sufficient to cause eutrophication; however, the contribution of atmospheric nitrogen deposition
to total nitrogen load is calculated for some estuaries and can be >40%  (REA 5.1.1).
       The evidence is sufficient to infer a causal relationship between N deposition and the
alteration of species richness, species composition and biodiversity in estuarine ecosystems (ISA
4.3.4.2 and 3.3.5.4).  Atmospheric and non-atmospheric sources of N contribute to increased
phytoplankton and algal productivity, leading to eutrophication.  Shifts in community
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composition, reduced hypolimnetic DO, decreases in biodiversity, and mortality of submerged
aquatic vegetation are associated with increased N deposition in estuarine systems.
3.2.2  Ecosystem sensitivity to nutrient enrichment
       The numerous ecosystem types that occur across the U.S. have a broad range of
sensitivity to N deposition (Clark and Tilman 2008; Aber et al, 2003; Fenn et al, 2003; Fenn et
al, 2007; Rueth et al., 2003; Egerton-Warburton and Allen 2000; Williams et al., 1996; and
additional studies summarized in ISA Table 4-4). Increased deposition to N-limited ecosystems
can lead to production increases that may be either beneficial  or adverse depending on the
system and management goals.
       Organisms in their natural environment are commonly adapted to a specific regime of
nutrient availability. Change in the availability of one important nutrient, such as N, may result
in an imbalance in ecological stoichiometry, with effects on ecosystem processes, structure and
function (Sterner and Elser, 2002). In general, N deposition to terrestrial ecosystems causes
accelerated growth rates in some species deemed desirable in  commercial forests but may lead to
altered competitive interactions among species and nutrient imbalances, ultimately affecting
biodiversity. The onset of these effects occurs with N deposition levels as low as 3 kg N/ha/yr in
sensitive terrestrial ecosystems to N deposition. In aquatic ecosystems, N that is both leached
from the soil and directly deposited to the water surface can pollute the surface water. This
causes alteration of the diatom community at levels as low as  1.5 kg N/ha/yr in sensitive
freshwater ecosystems.
       The degree of ecosystem effects lies at the intersection of N loading and N-sensitivity.
N-sensitivity is predominately driven by the degree to which growth is limited by nitrogen
availability.  Grasslands in the western United States are typically N-limited ecosystems
dominated by a diverse mix of perennial forbs and grass species (Clark and Tilman, 2008;
Suding et al., 2005). A meta-analysis by LeBauer and Treseder (2008) indicated that N
fertilization increased aboveground growth in all non-forest ecosystems except for deserts. In
other words, almost all terrestrial ecosystems are N-limited and will be altered by the addition of
anthropogenic nitrogen (LeBauer and Treseder, 2008). Likewise, a freshwater lake or stream
must be N-limited to be sensitive to N-mediated eutrophication. There are many examples of
fresh waters that are N-limited or N and phosphorous (P) co-limited (ISA 3.3.3.2). In a meta-
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analysis that included 653 datasets, Elser et al. (2007) found that N-limitation occurred as
frequently as P-limitation in freshwater ecosystems. Additional factors that govern the sensitivity
of ecosystems to nutrient enrichment from N deposition include rates and form of N deposition,
elevation, climate, species composition, plant growth rate, length of growing season, and soil N
retention capacity (ISA 4.3). Less is known about the extent and distribution of the terrestrial
ecosystems in the U.S. that are most sensitive to the effects of nutrient enrichment from
atmospheric N deposition compared to acidification.
       Because the productivity of estuarine and near shore marine ecosystems is generally
limited by the availability of N, they are susceptible to the eutrophication effect of N deposition
(ISA 4.3.4.1). A recent national assessment of eutrophic conditions in estuaries found the most
eutrophic estuaries were generally those that had large watershed-to-estuarine surface area, high
human population density, high rainfall and  runoff, low dilution, and low flushing rates (Bricker
et al., 2007).  In the REA, the National Oceanic and Atmospheric Administration's (NOAA)
National Estuarine Eutrophication Assessment (NEEA) assessment tool, Assessment of
Estuarine Tropic Status (ASSETS) categorical Eutrophication Index (El) (Bricker et al., 2007)
was used to evaluate  eutrophication due to atmospheric loading of nitrogen. ASSETS El is an
estimation of the likelihood that an estuary is experiencing eutrophication or will experience
eutrophication based  on five ecological indicators: chlorophyll a, macroalgae, dissolved oxygen,
nuisance/toxic algal blooms and submerged  aquatic vegetation (SAV) (Bricker et al., 2007).
       In the REA, two regions were selected  for case study analysis using ASSETS El, the
Chesapeake Bay and  Pamlico Sound.  Both regions received an ASSETS El rating of Bad
indicating that the estuary had moderate to high pressure due to overall human influence and  a
moderate high to high eutrophic condition (REA 5.2.4.1 and 5.2.4.2). These results were then
considered with SPAtially Referenced Regression (SPARROW) modeling to develop a response
curve to examine the  role of atmospheric nitrogen deposition in achieving a desired decrease  in
load.  To change the Neuse River Estuary's El score from Bad to Poor not only must 100% of
the total atmospheric  nitrogen deposition be  eliminated, but considerably more nitrogen from
other sources as well  must be controlled (REA section 5.2.7.2). In the Potomac River  estuary, a
78% decrease of total nitrogen could move the El score from Bad to Poor (REA  5.2.7.1).  The
results of this analysis indicated decreases in atmospheric deposition alone could not eliminate
coastal eutrophication problems due to multiple non-atmospheric nitrogen inputs (REA 7.3.3).
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However, by decreasing atmospheric contributions, it may help avoid the need for more costly
controls on nitrogen from other sources.  In addition, the somewhat arbitrary discreteness of the
El scale can mask the benefits of decreases in nitrogen between categories.
       In general, estuaries tend to be N-limited (Elser et al, 2008), and many currently receive
high levels of N input from human activities to cause eutrophication (Howarth et al., 1996;
Vitousek and Howarth, 1991). Atmospheric N loads to  estuaries in the U.S.  are estimated to
range from 2-8% for Guadalupe Bay, TX on the lowest end to as high as 72% for St Catherines-
Sapelo estuary, GA (Castro et al., 2003). The Chesapeake Bay  is an example of a large, well-
studied and severely eutrophic estuary that is calculated to receive as much as 30% of its total N
load from the atmosphere.
3.2.3  Magnitude of ecosystem responses
Terrestrial ecosystems
       Little is known about the full extent and distribution of the terrestrial ecosystems in the
U.S. that are most sensitive to impacts caused by nutrient enrichment from atmospheric N
deposition. As previously stated, most terrestrial ecosystems are N-limited, therefore they are
sensitive to perturbation caused by N additions (LeBauer and Treseder, 2008). Effects are most
likely to occur where areas of relatively high atmospheric N deposition intersect with N-limited
plant communities.  The alpine ecosystems of the Colorado Front Range, chaparral watersheds of
the Sierra Nevada, lichen and vascular plant communities in the San Bernardino Mountains and
the Pacific Northwest, and the southern California coastal sage scrub (CSS)  community are
among the most sensitive terrestrial ecosystems. There is growing evidence  that existing
grassland ecosystems in the western United States are being altered by elevated levels of N
inputs, including inputs from atmospheric deposition (Clark and Tilman, 2008; Suding et al.,
2005).
       In the eastern U.S., the degree of N saturation of the terrestrial ecosystem is often
assessed in terms of the degree of NOs  leaching from watershed soils into ground water or
surface water.  Stoddard (1994) estimated the number of surface waters at different stages of
saturation across several regions in the eastern U.S. Of the 85 northeastern watersheds examined
60% were in Stage 1 or Stage 2 of N saturation on a scale of 0 (background  or pretreatment) to 3
(visible decline). Of the northeastern sites for which adequate data were available for assessment,
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those in Stage 1 or 2 were most prevalent in the Adirondack and Catskill Mountains. Effects on
individual plant species have not been well studied in the U.S. More is known about the
sensitivity of particular plant communities. Based largely on results obtained in more extensive
studies conducted in Europe, it is expected that the more sensitive terrestrial ecosystems include
hardwood forests, alpine meadows, arid and semi-arid lands,  and grassland ecosystems (ISA
3.8.2).
       The REA used published research results (REA 5.3.1  and ISA Table 4.4) to identify
meaningful ecological benchmarks associated with different levels of atmospheric nitrogen
deposition. These are given by Figure 3-4.  The sensitive areas and ecological indicators
identified by the ISA were analyzed further in the REA to create a national map that illustrates
effects observed from ambient and experimental atmospheric nitrogen deposition loads in
relation to Community Multi-scale Air Quality (CMAQ) 2002 modeling results and NADP
monitoring data. This map, reproduced in Figure 3-5, depicts the sites where empirical effects of
terrestrial nutrient enrichment have been observed and site proximity to elevated atmospheric N
deposition.
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            5.2-
            5.3
Rocky Mountain alpine lakes: shift in diatom community dominance (Baron, 2006)


 •  Southern California: N growth requirement threshold (Wood et al., 2006)
 •  San Bernardino Mountains and Sierra Nevada Mountains: acidophytic lichen
    decline in MCF (Fenn et al., 2008)

 •   Eastern Rocky Mountain Slope: low carbon:nitrogen; low lignin:nitrogen (Baron et
    al., 2000)
 •   Eastern Rocky Mountain Slope: increased foliar nitrogen; increased mineralization
    (Baron etal., 2000)

   •  San Bernardino Mountains and Sierra Nevada Mountains: shift from acidophytic
      to neutral or nitrogen-tolerant lichen in MCF (Fenn et al., 2008)
   •  Minnesota grasslands: decreased plant species (Clark and Tilman, 2008)
              7-12
              9.8-10.2

                              Northeast U.S.: NO3 leaching (Aber et al., 2003)
          Bay Area, CA: Increased cover of nonnative grasses; decreased native
          grasses (Weiss, 1999)

          San Bernardino Mountains and Sierra Nevada Mountains: loss of acidophytic
          lichen in MCF (Fenn et al., 2008)
          Southern California: shift in mycorrhizal species in CSS (Egerton-Warburton
          and Allen, 2000)
          Southern California: shift from native species to invasive grasses in CSS (Allen,
          2008)*
           •  San Bernardino Mountains: high dissolved organic nitrogen (Meixner and
              Fenn, 2004)
           •  San Bernardino Mountains: nitrogen saturation (Meixner and Fenn, 2004)
                                     Increased nitrogen in lichen (Fenn et al., 2007)
                                                MCF: NO3 leaching (Fenn et al., 2008)
                                                MCF: 25% decrease in fine-root biomass (Fenn et al., 2008)

                                                 Southern California: NO3"  leaching (Fenn et al., 2003)
                                                 Southern California: high foliar nitrogen (Bytnerowicz and
                                                 Fenn, 1996)
                                                 Los Angeles Basin, California: High NO emissions
                                                 (Bytnerowicz and Fenn, 1996)
                                                                   Fraser Experimental Forest, CO:
                                                                   increased foliar nitrogen; increased
                                                                   mineralization (Rueth et al., 2003)**
                          10
             15       20        25       30        35       40
             Nitrogen Deposition, kg/ha/yr
45
         " Personal communication, 2008, Abo referenced in Bobbink et, al .2010, Ecological Application5,20(1);30-59 and USDS FS, 2010,
         http:/,'www.nfs.fs.fcd.ua'cl6an_air_water^ei&an_wa[ar/critlcal_lDads/local-fesotjrc6s.'docs/Emplrical_CLS_of_N_100414.pdf
         "Nitrogen deposition levels include ambient and experimental additions.
Figure 3-4.   Benchmarks of atmospheric nitrogen deposition for several ecosystem indicators
(REA 5.3.1.2) MCF-Mixed Conifer Forest, CSS-Coastal Sage Scrub.
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                         Total N Deposition
                         (kg/hafyr)
                         •• High: 66507

                          _ Low 0,761
 Legend

fj Effects from Observational Research

/ \ Effects from Experimental Research
  I National Parks
  I National Forests
                                Effects from Observational Research [Ambient N Deposition only)
                                                                                            Effects from Experimental Research fAmbient H deposition + Experimental N Additions)
                        1, Impacts on lichen communities
                        (CA mixed conifer forests: 3.1 kg/hatyr;
                        Columbia R  Gorge, OR/VW: 11.5-25.1)

                        2. Evidence or threatened and endangered
                        species impacted (San Francisco Bay, CA:
                        10-15k»*Tia/yr)

                        3 N saturation, high nil-ate in s,treanwater. soil,
                        leaves; high NO emissions (LA. CAAir Basin:
                        25-45 kgflna/yr; northeast US: 7-tQ kgmaVyr)

                        4. N saturation, high nitrate in streamwater
                        (San Bernardino Mtns.. CA:  18.8 kg/ha/yr)

                        5, N saturation, high Dissolved Inorganic N
                        (San Bernardino Mtns., CA:  11-40 kg.'ha/yr)
               6 Increased free mortality and beetle activity
               (San Bernardino Mtns.. CA: 82 kg/na/yO

               7. N enrichment o' eutrophication of lakes
               (Loch Vale- CO: 0.5-1.5 kg'ha'yr
               Niwol Ridge, CO: 471-165 kgftia/yr)

               8. Increased soil and foliar N concentration
               (Eastern slope of Rocky Mtns.:
               3.5-3 6 kortia/yr)

               9. Decrease in pilcher plant (Hawley Bog, MA
               and Molly Bog, VA~  10-14 kg'ha'yr)

               10. Nitrate leaching (Adirondack lakes,
               New England: 8-10 kgrha/yr)
1 1 Decreased diversity of mycorrhizal cornmu nities
(southern California: -10 kg/ha'yr Northern
Michigan:
12. Decreased desert creosote bush,
increased non native grasses
(Mojave Desert, CA: 46-62 kg/hafyr )

13 Community structure changes
(Chihuahuan Desert, CA: 21.7 kg/hafyrand up)

14. Alpine meadows' elevated nitrate tevets
in runoff (Colo Front Range: 20 kg.'ha/yr)

15. Alpine meadows' shift toward hairgrass
(Niwot Ridge. CO:
16. Increased soil and foliar N concentration
{Raser Forest, CO: 28.2-30.5 kgrtia/yr)

17. Increased N mineralization rates and
nitnficalion (Loch Vale. CO: 26.7 kg/ha/yr)

18, Loss of grassland species (Cedar Creek, MN:
16 kg/ha'yrand up*)

19. Enhanced growth of black cherry and
yellow poplar possible decline in red maple
vigor; increased foliar N (Fernow Forest, WV
52.5 kgrtia/yr)

'The authors estimated lhat Ihe critical toad for chronic
N dflposilion eKi&ts at 5.3 kg.'ha.'yr us^ng mcxjels,
no.vevc' fro observed effects woro produced as a
rwuH of chronic additions grcatcf Ihon 10 kg NJh»fyr,
Figure 3-5.  Observed effects  from  ambient and experimental atmospheric nitrogen deposition loads in relation to using CMAQ 2002
modeling results and NADP monitoring data.
Citations for effect results are from the ISA, Table 4.4 (U.S. EPA, 2008) 1= Fenn et. al. (2008), 2=Weiss (1999), 3=Bytnerowicz and Fenn (1996), 4=Fenn et al. (2000), 5=
Meixner and Fenn (2004), 6=Jones et al. 2004, 7=Baron (2006), 8=Baron et al. (2000), 9=Gotelli and Ellison (2002), 10=Stoddard et al. (1994), 11 =Egerton Warburton and Allen
(2000), 12=Brooks (2003), 13=Baez et al. (2007), 14=Bowman et al. (2006), 15=Bowman et al. (1995), 16=Rueth et al. (2003), 17=DeWalle et al. (2006),  18=Clark and Tillman
(2008), 19=Rueth et al. 2003
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       Based on information in the ISA and initial analysis in the REA, further case
study analyses on terrestrial nutrient enrichment of ecosystems were developed for the
CCS community and Mixed Conifer Forest (MCF) (EPA, 2009). Geographic
information systems (GIS) analysis supported a qualitative review of past field research
to identify ecological benchmarks associated with CSS and mycorrhizal communities, as
well as MCF's nutrient-sensitive acidophyte lichen communities, fine-root biomass in
Ponderosa pine, and leached nitrate in receiving waters.
       The ecological benchmarks that were identified for the CSS and the MCF are
included in the suite of benchmarks identified in the ISA (ISA 3.3). There are sufficient
data to confidently relate the ecological effect to a loading of atmospheric nitrogen. For
the CSS community, the following ecological benchmarks were identified:
    •  3.3 kg N/ha/yr - the amount of nitrogen uptake by a vigorous stand of CSS; above
       this level, nitrogen may no longer be limiting
    •  10 kg N/ha/yr - mycorrhizal community changes
For the MCF community, the following ecological benchmarks were identified:
    •  3.1 kg N/ha/yr - shift  from  sensitive to tolerant lichen species
    •  5.2 kg N/ha/yr - dominance of the tolerant lichen species
    •  10.2 kg N/ha/yr - loss of sensitive lichen species
    •  17 kg N/ha/yr - leaching of nitrate into streams.
       These benchmarks, ranging from 3.1 to 17 kg N/ha/yr, were compared to 2002
CMAQ/NADP data to discern any  associations between atmospheric deposition and
changing communities. Evidence supports the finding that nitrogen alters CSS and MCF.
Key findings include the following: 2002 CMAQ/NADP nitrogen deposition data show
that the 3.3 kg N/ha/yr benchmark has been exceeded in more than 93% of CSS areas
(654,048 ha). These deposition levels are a driving force in the degradation of CSS
communities. Although CSS decline has been observed in the absence of fire, the
contributions of deposition and fire to the CSS decline require further research. CSS is
fragmented into many small parcels, and the 2002 CMAQ/NADP 12-km grid data are not
fine enough to fully validate the relationship between CSS distribution, nitrogen
deposition,  and fire. 2002 CMAQ/NADP nitrogen deposition data exceeds the 3.1 kg
N/ha/yr benchmark in more than 38% (1,099,133 ha) of MCF areas, and nitrate leaching
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has been observed in surface waters. Ozone effects confound nitrogen effects on MCF
acidophyte lichen, and the interrelationship between fire and nitrogen cycling requires
additional research.
Freshwater ecosystems
        The magnitude of ecosystem response may be thought of on two time scales,
current conditions and how ecosystems have been altered since the onset of
anthropogenic N deposition.  As noted previously, Elser et al.  (2007) found that N-
limitation occurs as frequently as P-limitation in freshwater ecosystems (ISA 3.3.3.2).
Recently, a comprehensive study of available data from the northern hemisphere surveys
of lakes along gradients of N deposition show increased  inorganic N concentration and
productivity to be correlated with atmospheric N deposition (Bergstrom and Jansson
2006). The results are unequivocal evidence of N limitation in lakes with low ambient
inputs of N, and increased N concentrations in lakes receiving N solely from atmospheric
N deposition (Bergstrom and Jansson, 2006). These authors suggested that most  lakes in
the  northern hemisphere may have originally been N-limited,  and that atmospheric N
deposition has changed the balance of N and P in lakes.
       Available  data suggest that the increases in total N deposition do not have to be
large to elicit an ecological effect. For example, a hindcasting exercise determined that
the  change in Rocky Mountain National Park lake algae  that occurred between 1850 and
1964 was associated with an increase in wet N deposition that was only about 1.5 kg
N/ha (Baron, 2006). Similar changes inferred from lake sediment cores of the Beartooth
Mountains of Wyoming also occurred at about 1.5 kg N/ha deposition (Saros et al.,
2003). Pre-industrial inorganic N deposition is estimated to have been only 0.1 to 0.7 kg
N/ha based on measurements from remote parts of the world (Galloway et al., 1995;
Holland et al., 1999). In the western U.S., pre-industrial, or background, inorganic N
deposition was estimated by (Holland et al., 1999) to range from 0.4 to 0.7 kg N/ha/yr.
       Eutrophication effects from N deposition are most likely to be manifested in
undisturbed, low nutrient  surface waters such as those found in the higher elevation areas
of the western U.S. The most severe eutrophication from N deposition effects is expected
downwind of major urban and agricultural centers.  High concentrations of lake or
streamwater NOs  , indicative of ecosystem saturation, have been found at a variety of
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locations throughout the U.S., including the San Bernardino and San Gabriel Mountains
within the Los Angeles Air Basin (Fenn et al., 1996), the Front Range of Colorado
(Baron et al., 1994; Williams et al., 1996), the Allegheny mountains of West Virginia
(Gilliam et al., 1996), the Catskill Mountains of New York (Murdoch and Stoddard,
1992; Stoddard, 1994), the Adirondack Mountains of New York (Wigington et al., 1996),
and the Great Smoky Mountains in Tennessee (Cook et al., 1994) (ISA 3.3.8).
Estuaries
       In contrast to terrestrial and freshwater systems, atmospheric N load to estuaries
contributes to the total load but does not necessarily drive the effects since other
combined sources of N often greatly exceed N deposition.  In estuaries, N-loading from
multiple anthropogenic and non-anthropogenic pathways leads to water quality
deterioration, resulting in numerous effects including hypoxic zones, species mortality,
changes in community composition and harmful algal blooms that are indicative of
eutrophication. The following summary is a concise overview of the known or
anticipated effects of nitrogen enrichment on estuaries within the United  States.
       There is a scientific consensus that nitrogen-driven eutrophication in shallow
estuaries has increased over the past several decades and that the environmental
degradation of coastal ecosystems due to nitrogen, phosphorus, and other inputs  is now a
widespread occurrence (Paerl et al., 2001). For example, the frequency of phytoplankton
blooms and the extent and severity of hypoxia have increased in the Chesapeake Bay
(Officer et al., 1984) and Pamlico estuaries in North Carolina (Paerl et al., 1998) and
along the continental shelf adjacent to the Mississippi  and Atchafalaya rivers' discharges
to the Gulf of Mexico  (Eadie et al., 1994).
       A recent national assessment of eutrophic  conditions in estuaries found that 65%
of the assessed systems had moderate to high overall eutrophic conditions and generally
received the greatest N loads from all sources, including atmospheric and land-based
sources (Bricker et al., 2007). Most eutrophic estuaries occurred in the mid-Atlantic
region and the estuaries with the lowest degree of eutrophication were in the North
Atlantic (Bricker et al., 2007). Other regions had mixtures  of low, moderate, and high
degrees of eutrophication (ISA 4.3.4.3).
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        The mid-Atlantic region is the most heavily impacted area in terms of moderate or
 high loss of submerged aquatic vegetation due to eutrophication (ISA 4.3.4.2).
 Submerged aquatic vegetation is important to the quality of estuarine ecosystem habitats
 because it provides habitat for a variety of aquatic organisms, absorbs excess nutrients,
 and traps sediments (ISA 4.3.4.2). It is partly because many estuaries and near-coastal
 marine waters are degraded by nutrient enrichment that they are highly sensitive to
 potential negative impacts from nitrogen addition from atmospheric deposition.
 3.2.4   Key uncertainties associated with nutrient enrichment
        There are different levels of uncertainty associated with relationships between
 deposition,  ecological effects and ecological indicators. The criteria used in the REA to
 evaluate the degree of confidence in the data, modeling and ecological effect function are
 detailed in Chapter 7 of the REA and summarized in section 3.1.4 of this chapter.
Aquatic ecosystems
        The approach for assessing atmospheric contributions to total nitrogen loading in
 the REA, was to consider the main-stem river to an estuary (including the estuary) rather
 than an entire estuary system or bay.  The biological indicators used in the NOAA
 ASSETS El required the evaluation of many national databases including the US
 Geological  Survey National Water Quality Assessment (NAWQA) files, EPA's
 STORage and RETrieval (STORET) database, NOAA's Estuarine Drainage Areas data,
 and EPA's water quality standards nutrient criteria for rivers and lakes (REA Appendix 6,
 Table 1.2.-1). Both the SPARROW modeling for nitrogen loads and assessment of
 estuary conditions under NOAA ASSETS El, have been applied on a national scale.  The
 REA concludes that the available data are medium quality with intermediate confidence
 about the use of these data and their values for extrapolating to a larger regional area
 (REA 7.3.1). Intermediate confidence is associated with the modeling approach using
 ASSETS El and SPARROW.  The REA states there is low confidence with the
 ecological effect function due to the results of the analysis which indicated that
 reductions in atmospheric deposition alone could not solve coastal eutrophication
 problems due to multiple non-atmospheric nitrogen inputs (REA 7.3.3).
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Terrestrial ecosystems
       Ecological thresholds are identified for CSS and MCF and these data are
considered to be of high quality, however, the ability to extrapolate these data to larger
regional areas is limited (REA 7.4.1). No quantitative modeling was conducted or
ecological effect function developed for terrestrial nutrient enrichment reflecting the
uncertainties associated with these depositional effects.

3.3    ECOLOGICAL EFFECTS ASSOCIATED WITH GAS-PHASE OXIDES
       OF NITROGEN AND SULFUR
       Acidifying deposition and nitrogen enrichment are the main focus of this policy
assessment;  however, there are other known ecological effects are attributed to gas-phase
oxides of nitrogen and sulfur. Acute and chronic exposures to gaseous pollutants such as
sulfur dioxide (802), nitrogen dioxide (NO2), nitric oxide (NO), nitric acid (HNOs) and
peroxyacetyl nitrite (PAN) are associated with negative impacts to vegetation. The
current secondary NAAQS were set to protect against direct damage to vegetation by
exposure to  gas-phase NOX or SOX, such as foliar injury, decreased photosynthesis, and
decreased growth.  The following summary is a concise overview of the known or
anticipated effects to vegetation caused by gas phase N and S. Most phototoxic effects
associated with gas phase oxides of nitrogen and sulfur occur at levels well above
ambient concentrations observed in the U.S. (ISA 3.4.2.4).
3.3.1   Nature  of ecosystem responses to gas-phase nitrogen and sulfur
       The 2008 ISA found that gas phase N and S  are associated with direct phytotoxic
effects (ISA 4.4). The evidence is sufficient to infer a causal relationship between
exposure to  SO2 and injury to vegetation (ISA 4.4.1 and 3.4.2.1). Acute foliar injury to
vegetation from 862 may occur at levels above the current secondary standard (3-h
average of 0.50 ppm). Effects on growth, reduced photosynthesis and decreased yield of
vegetation are also associated with increased  862 exposure concentration and time of
exposure.
        The  evidence is sufficient to infer a causal relationship between exposure to NO,
NO2 and PAN and injury to vegetation (ISA 4.4.2 and 3.4.2.2). At sufficient
concentrations,  NO, NO2 and PAN can decrease photosynthesis and induce visible foliar
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injury to plants.  Evidence is also sufficient to infer a causal relationship between
exposure to HNOs and changes to vegetation (ISA 4.4.3 and 3.4.2.3).  Phytotoxic effects
of this pollutant include damage to the leaf cuticle in vascular plants and disappearance of
some sensitive lichen species.
3.3.2  Ecosystem sensitivity
       Vegetation in ecosystems near sources of gaseous oxides of nitrogen and sulfur or
where ambient concentrations of SO2, NO, NO2, PAN and HNOs are higher are more
likely to be impacted by these pollutants. Uptake of these pollutants in a plant canopy is a
complex process involving adsorption to surfaces (leaves, stems and soil) and absorption
into leaves (ISA 3.4.2). The functional relationship between ambient concentrations of
gas phase oxides of nitrogen and sulfur and specific plant response are impacted by
internal factors such as rate of stomatal conductance and plant detoxification
mechanisms, and external factors including plant water status, light, temperature,
humidity,  and pollutant exposure regime (ISA 3.4.2).
       Entry of gases into a leaf is dependent upon physical and chemical processes of
gas phase as well as to stomatal aperture.  The aperture of the stomata is controlled
largely by the prevailing environmental conditions, such as water availability, humidity,
temperature, and light intensity. When the stomata are closed, resistance to gas uptake is
high and the plant has a very low degree of susceptibility to injury. Mosses  and lichens
do not have a protective cuticle barrier to gaseous pollutants or stomata and are generally
more sensitive to gaseous sulfur and nitrogen than vascular plants (ISA 3.4.2).
       The appearance of foliar injury can vary significantly across species and growth
conditions affecting stomatal conductance in vascular plants (REA 6.4.1). For example,
damage to lichens from SO2 exposure include decreases in photosynthesis and
respiration, damage to the algal component of the lichen, leakage of electrolytes,
inhibition of nitrogen fixation, decreased K+ absorption, and structural changes (Belnap
et al, 1993; Farmer et al, 1992, Hutchinson et  al, 1996).
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3.3.3  Magnitude of ecosystem responses to gas-phase effects of oxides of nitrogen
       and sulfur
       The phytotoxic effects of gas phase oxides of nitrogen and sulfur are dependent
on the exposure concentration and duration and species sensitivity to these pollutants.
Effects to vegetation associated with oxides of nitrogen and sulfur, are therefore, variable
across the U.S. and tend to be higher near sources of photochemical smog. For example,
SC>2 is considered to be the primary factor contributing to the death of lichens in many
urban and industrial areas, with fruticose lichens being more susceptible to SCh than
many foliose and crustose species (Hutchinson et al, 1996).
       The ISA states there is very limited new research on phytotoxic effects of NO,
NC>2, PAN and HNOs  at concentrations currently observed in the United States with the
exception of some lichen species (ISA 4.4). Past and current HNOs concentrations may
be contributing to the decline in lichen species in the Los Angeles basin (Boonpragob and
Nash 1991; Nash and Sigal, 1999; Riddell et al., 2008). PAN is a very small component
of nitrogen deposition in most areas of the United States (REA 6.4.2).  Current deposition
of HNOs is contributing to N saturation of some ecosystems  close to sources of
photochemical smog (Fenn et al., 1998) such as the MCF's of the Los Angeles basin
mountain (Bytnerowicz et al., 1999).  Most phototoxic effects associated with gas phase
oxides of nitrogen and sulfur occur at levels well above ambient concentrations observed
in the U.S. (ISA 3.4.2.4).

3.4    SUMMARY
       In summary, oxides of nitrogen and sulfur in the atmosphere contribute to effects
on individual species and ecosystems through direct contact with vegetation, and more
significantly through deposition to sensitive ecosystems.  The ISA concludes that the
evidence is sufficient to  conclude causal relationships between acidifying deposition of N
and S and effects on freshwater aquatic ecosystems and terrestrial ecosystems, and
between nitrogen nutrient enrichment and effects on sensitive terrestrial and freshwater
aquatic ecosystems. The ISA also concludes that a causal relationship is supported
between nitrogen nutrient enrichment and effects on estuarine ecosystems; however, the
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contribution of atmospheric oxidized nitrogen relative to reduced nitrogen and non-
atmospheric nitrogen is more difficult to determine.
       The REA provides additional support that under recent conditions; deposition
levels have exceeded benchmarks for ecological indicators of acidification and nutrient
enrichment that indicate that effects are likely to be widespread in lakes and streams
within sensitive ecosystems.
       When considering all of the depositional effects together, it is clear that more
research has been done on quantifying the relationship between atmospheric deposition
and aquatic acidification than on terrestrial acidification or nitrogen enrichment of
terrestrial or aquatic ecosystems.  Therefore, staff concludes that it is appropriate to focus
on the development of a secondary NAAQS to protect against the deposition-related
effects of aquatic acidification in this review. Chapter 5 addresses the co-benefits that
such a standard would  afford with regard to protection against other deposition-related
effects.
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Nash TH; Sigal LL. 1999. Epiphytic lichens in the San Bernardino mountains in relation
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Norton S; Kahl J; Fernandez I. 1999. Altered soil-soil water interactions inferred from
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Officer CB; Biggs RB; Taft JL; Cronin LE; Tyler MA; Boynton WR. 1984. Chesapeake
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Paerl HW; Boynton WR; Dennis RL; Driscoll CT; Greening HS; Kremer JN; Rabalais
       NN; Seitzinger SP. 2001. Atmospheric deposition of nitrogen in coastal waters:
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       loading in coastal water bodies: an atmospheric perspective. (Coastal and
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       Union.
Paerl H; Pinckney J; Fear J; Peierls B.  1998. Ecosystem responses to internal and
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Parker KE. 1988. Common loon reproduction and chick feeding on acidified lakes  in the
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Parker DR; Zelazny LW; Kinraide TB. 1989. Chemical speciation and plant toxicity of
       aqueous aluminum. In: Lewis TE (Ed.), Environmental chemistry and toxicology
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Peterjohn WT; Adams MB; Gilliam FS. 1996. Symptoms of nitrogen saturation in two
       central Appalachian hardwood forest ecosystems.Biogeochemistry, 35, 507-522.
Reuss JO. 1983. Implications of the calcium-aluminum exchange system for the effect of
       acid precipitation on soils. J Environ Qual, 12, 591-595.
Riddell J; Nash lii TH; Padgett P. 2008. The effect of HNO3 gas on the lichen Ramalina
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Rueth, H.M., J.S. Baron, and EJ. Allstott. 2003. Responses of Engelmann spruce forests
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Saros JE; Interlandi SJ; Wolfe AP; Engstrom DR. 2003. Recent changes in the diatom
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      Academic Press.
Schwinning S; Starr BI; Wojcik NJ; Miller ME; Ehleringer JE; Sanford RL Jr. 2005.
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          4  CONSIDERATIONS OF ADVERSITY TO PUBLIC WELFARE

4.1    INTRODUCTION
       Characterizing a known or anticipated adverse effect to public welfare is an important
component of developing any secondary NAAQS. According to the Clean Air Act, welfare
effects include:
       effects on soils, water, crops, vegetation, manmade materials, animals, wildlife,
       weather, visibility, and climate, damage to and deterioration of property, and
       hazards to transportation, as well as effect on economic values and on personal
       comfort and well-being, whether caused by transformation, conversion, or
       combination with  other air pollutants (CAA,  Section 302(h)).

While the text above lists a number of welfare effects, these effects do not define public
welfare in and of themselves.
       Although there is no specific definition of adversity to public welfare, the paradigm of
adversity to public welfare as  deriving from disruptions in ecosystem structure and function has
been used broadly by EPA to categorize effects of pollutants from the cellular to the ecosystem
level. An evaluation of adversity to public welfare might consider the likelihood, type,
magnitude, and spatial scale of the effect as well as the potential for recovery and any
uncertainties relating to these considerations.
       Similar concepts were used in past reviews of secondary NAAQS for ozone, PM (relating
to visibility), as well as in initial reviews of effects from lead deposition.  Because oxides of
nitrogen and sulfur are  deposited from ambient sources  into ecosystems where they affect
changes to organisms, populations and ecosystems, the concept of adversity to public welfare as
related to impacts on the public from alterations in structure and function of ecosystems is
appropriate for this review.  Other information that may be helpful to consider includes the role
of critical  loads and ecosystem service impacts as benchmarks or  measures of impacts on
ecosystems that may affect public welfare.  Ecosystem services can be related directly to
concepts of public welfare to inform discussions of societal adverse impacts.
       Subsequent sections of Chapter 4 discuss benchmarks of adversity from other EPA
programs, other federal agencies and the European Union. We will also define and discuss
ecosystem services and the role of economics in defining  adversity to public welfare. Finally we
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will discuss the results of analyses relating adversity to public welfare to aquatic acidification,
terrestrial acidification, aquatic nutrient enrichment, and terrestrial nutrient enrichment.

4.1.1   Benchmarks from Other EPA Programs
       Various federal laws and policies exist to protect ecosystem health. How other EPA
programs and offices consider ecosystem effects in carrying out their programs can help inform
the Administrator when she evaluates the adversity of ecosystem impacts on public welfare.
From the 1996 National Acid Precipitation Assessment Program Report to Congress: " The 1990
Clean Air Act Amendments require that the National Acid Precipitation Assessment Program
(NAPAP) prepare biennial reports to Congress, and that "every four years ... the report ...  shall
include the reduction in deposition rates that must be achieved in order to prevent adverse
ecological effects" (Public Law 101-549, Title IX, Section 903 (j)(3)(F)(i), codified as amended
at 42 USC §7403(j)(3)(F)(I)). Although the term adverse ecological effects is not specifically
defined in the Clean Air Act Amendments, a working definition can be derived from relevant
statements at various locations in the statute. Congress expresses its concern with ecological
components (the scope is broad and inclusive, since ecology encompasses the interrelationships
of organisms and their environment) in the preceding subsection (E) of the statute. That
subsection requires reporting on "the status of ecosystems (including forest and surface waters)
... affected by acid deposition ... including changes in surface water quality  and forest and soil
conditions ... [and] high elevation watersheds" (42 USC §7403(j)(3)(E)(iii-v)). The adverse
effects of concern to Congress, as evidenced in its findings and declaration of purpose, are the
"dangers to the  public health and welfare ... including injury ... damage ... and ... deterioration"
(42 USC §7401 (a)). Based on the intent of Congress, as expressed above and elsewhere in the
Clean Air Act, and shaped by indications of intent expressed in other relevant environmental
statutes and regulations, the following working definition of adverse ecological effects has  been
derived and is used in the preparation of the NAPAP report:
       any injury (i.e., loss of chemical or physical quality or viability) to any
       ecological or ecosystem component, up to and including at the regional
       level, over both long and short terms. Similarly, adverse effects for other
       areas of concern addressed in this report—i.e., visibility, materials, and
       human health—consist of loss of quality up to and including at the
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       regional level, over both long and short terms."
       As  another  example, an  effect may be  considered adverse to public welfare  if it
contributes to the inability of areas to meet water quality objectives as defined by the Clean
Water Act. The following federal statutes and policies may prove helpful to consider.
Ozone NAAQS Review
       The evaluation of adversity from a public welfare perspective in the context of ozone and
particulate matter (PM) are relevant to this current review.  Both ozone and PM have
documented effects on ecological receptors. These criteria pollutants are being reviewed on a
schedule as part of the NAAQS process. The ozone secondary standard is currently under
reconsideration from the 2008 ruling with a proposal was published January 6, 2010.  The final
Policy Assessment for PM is being developed and is expected to be finalized in the fall of 2010.
       For the purposes of the reconsideration of the secondary standard for ozone , welfare
effects of ozone are primarily limited to vegetation. These effects begin at the level of the
individual cell and accumulate up to the level of whole leaves and plants. If effects occur on
enough individual plants within the population, communities and ecosystems may be impacted.
Prior to the 2008 ozone review, Ozone vegetation effects were classified as either "injury" or
"damage" (FR 72 37889). "Injury" was  defined as; encompassing all plant reactions,  including
reversible changes or changes in plant metabolism, quality or reduced growth that does not
impair the intended use of the plant while "damage" includes those injury effects that reach
sufficient magnitude  as to decrease or impair the intended use of the plant (FR 72 37890). The
"intended use" of the plant was imbedded with the concept of adversity to public welfare.
Ozone-associated "damage" was considered adverse if the intended use of the plant was
compromised (i.e. crops, ornamentals, plants located in Class I areas).  Effects of ozone on
single plants or species grown in monocultures such as agricultural crops and managed forests
were evaluated without consideration of potential effects on natural forests or entire ecosystems.
       In the 2008 rulemaking, EPA expanded the characterization of adversity beyond the
individual plant level and this language  is continued in the 2010 ozone reconsideration. The
2008 final rule and 2010 proposal conclude that a determination of what constitutes an "adverse"
welfare effect in the context of secondary NAAQS review can appropriately occur by
considering effects at higher ecological levels (populations, communities, ecosystems) as
supported by recent literature. The ozone review uses the example of the construct presented in
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Hogsett et al. (1997) as a model for assessing risks to forests.  This study suggests that adverse

effects could be classified into one or more of the following categories: (1) economic production,

(2) ecological structure, (3) genetic resources, and (4) cultural values.  Another recent

publication,  "A Framework for Assessing and Reporting on Ecological Condition: an SAB

report" (Young and Sanzone, 2002) provides additional support for expanding the consideration

of adversity  beyond the species level and at higher levels by making explicit the linkages

between stress-related effects at the species level and at higher levels within an ecosystem

hierarchy (See Figure 4-1).
               Hydrologic liberation
               Habitat conversion
               Habitat fragmentation
               Climate change
               Invasive non-native species
               Turbidity/sedimentation
               Pesticides
               Diiejse/pest outbreaks
               Nutrientpulses
               Mftals
               Dissolved oxygen depletion
               Ozone (tropospheric)
                                            Hydrologic alteration
                                            Habita t conversion
                                            Habita t fragment* tion
                                            Cl»nate change
                                            Over-harvesting of vegetation
                                            Large-scale invasive
                                             species introductions
                                            Large-scale ctiseasc/pes t outbreaks
               Hydrohgic alteration
               Habitat conversion
               Climate change
               Turbidity/sedtmentation
               Pesticides
               Nutrient pulses
               Metals
               Dissolved oxygen depletion
               Ozone (tropospheric)
               Nitrogen oxides
                                         Chemical/
                                          Physical
                                        T
                                                              Natural
                                                            Disturbance
                                                            Hydrology/
                                                          Geomorphology.
                                                                             Hydrologic alteration
                                                                               Habitat conversion
                                                                                  Climate change
                                                                       Over-harvesting of 'vegetation
                                                                             Disease/pest outbreaks
                                                                               Altered ftre regime
                                                                              Altered flood regime
Ecological
Processes
                                                      t
  Hydrologic alteration
    Habitat conversion
 Habitat fragmentation
       Clirtiate change
Turbidity/sedimentation
                                            Hydrologic alteration
                                            Habitat conversion
                                            Clitnate change
                                            Pesticides
                                            Disease/pest outbreaks
                                            Nutrient pulses
                                            Dissolved oxygen depletion
                                            Nitrogen oxides


Figure 4-1.    Common anthropogenic stressors and the essential ecological attributes they
affect.  Modified from Young and Sanzone (2002).
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       In the 2008 ozone NAAQS review and current ozone NAAQS proposal, the
interpretation of what constitutes an adverse effect on public welfare can vary depending on the
location and intended use of the plant. The degree to which Os-related effects are considered
adverse to public welfare depends on the intended use of the vegetation and its significance to
public welfare (73 FR 16496). Therefore, effects on vegetation (e.g., biomass loss, foliar injury,
impairment of intended use) may be judged to have a different degree of impact on public
welfare depending, for example, on whether that effect occurs in a Class I area, a city park,
commercial cropland or private land.
       In the proposed ozone reconsideration in 2010 the Administrator has found that the types
of information most useful in informing the selection of an appropriate range of protective levels
is appropriately focused on information regarding exposures and responses of sensitive trees and
other native species that occur in protected areas such as Class I areas or on lands set aside by
States, Tribes and public interest groups to provide similar benefits to the public welfare. She
further notes that while direct links between O3  induced visible foliar injury symptoms and other
adverse effects (e.g., biomass loss) are not always found, visible foliar injury in itself is
considered by the National Park Service (NFS) to affect adversely air quality related values
(AQRV) in Class I areas, while the Administrator recognizes that uncertainty remains as to what
level of annual tree seedling biomass loss when compounded over multiple years should be
judged adverse to the public welfare, she believes that the  potential for such anticipated effects
should be considered in judging to what degree a standard should be precautionary (73 FR
16496). The range of proposed levels from 7-15 ppb includes at the maximum level of 15 ppb
protection of approximately 75% of seedlings from more than 10% biomass loss.
Prevention of Significant Deterioration Program
       The Clean Air Act's Prevention of Significant Deterioration (PSD) program (42 U.S.C.
7470) purposes include to "preserve, protect and enhance the air quality in national parks,
wilderness areas and other areas of natural, recreational, scenic or historic value . . . ." Also, the
PSD program charges the Federal Land Managers, including the NFS, with ". . . an affirmative
responsibility to protect the air quality related values . . . "within federal Class I lands. (42 U.S.C.
7475(d)(2)(B)).
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EPA Office of Water
       Section 101 of the Clean Water Act (CWA) (Declaration of Goals and Policy) states that
the objective of the CWA is to restore and maintain the chemical, physical, and biological
integrity of the Nation's waters and to attain, where possible, water quality that protects fish,
shellfish, wildlife and provides for water-based recreation.
       The CWA also authorizes EPA to develop water quality criteria as a guide for the states
to set water quality standards to protect aquatic life. In consideration of acidification effects,
EPA's Redbook, Quality Criteria for Water, published originally in 1976, recommends that
alkalinity be 20 mg/l\L  or more as CaCOs for freshwater aquatic life except where natural
concentrations are less.
Chesapeake Bay Total Maximum Daily Loads
       Under section 303(d) of the Clean Water Act, states, territories, and authorized tribes are
required to develop lists of impaired waters. These are waters that are too polluted or otherwise
degraded to meet the water quality standards set  by states, territories, or authorized tribes. The
law requires that these jurisdictions establish priority rankings for waters on the lists and develop
TMDLs for these waters. A Total Maximum Daily Load, or TMDL, is  a calculation of the
maximum amount of a pollutant that a waterbody can receive and still safely meet water quality
standards.  EPA is developing a TMDL for the Chesapeake Bay and its tributaries.  The
Chesapeake Bay Program has modeled the level  of nitrogen that can reach the Bay and still meet
the Bay's water quality standards.  The TMDL, with full public participation, will set waste load
allocations for point source discharges and load allocations for nonpoint sources of nitrogen.  Air
deposition to the Bay and its watershed, as a source category, will have a specific allocation.
        According to an EPA draft report responding to Section 202a of Executive Order 13508
(EPA,  2009a), within the Chesapeake Bay watershed, inorganic forms of nitrogen deposition
have been modeled and monitored. Organic forms have not been well quantified. Of the
inorganic nitrogen deposited from the air to the Chesapeake Bay watershed in 2002,
approximately 67 percent is oxidized nitrogen due to air emissions of NOX. The remaining 33
percent is in the form of reduced nitrogen from emissions of ammonia. There still remains
significant  uncertainty in the ammonia emissions inventory, which will be improved with further
emission and ambient measurements.
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       In 2002, about 87 million pounds (19 percent) of nitrogen load deposited on the
watershed was delivered to the Bay. An additional 22 million pounds of nitrogen were
atmospherically deposited  directly onto the surface of the tidal Bay's waters.
       Ammonia emissions, in 2002, were estimated to contribute approximately 147 of the 452
million pounds of nitrogen atmospheric deposition to the Bay watershed. About 80 percent of the
deposited ammonia loads were estimated to originate from agricultural operations and 20 percent
were from mobile and industrial sources, fires, and other sources.
       The allocation can  be used to calculate the level of ambient air concentrations of reactive
nitrogen that are likely to meet the deposition allocation.  To find the NOX portion of the
allocation one would subtract the reduced forms from the total allocation. If the total load to the
Bay of nitrogen from all the allocated source categories remains below the allocations, then the
Bay is expected to meet the water quality standards, which are set to protect the designated uses
of the Bay.  Since the designated uses are set by the states with public input, not meeting the
designated uses can be seen as having an adverse effect to public welfare.

4.1.2   Other Federal Agencies and the European Union
National Park Service
       The National Park  Service (NFS) is responsible for the protection of all resources within
the national  park system. These resources include those that are related to and/or dependent
upon good air quality, such as whole ecosystems and ecosystem components. The NFS, in its
Organic Act (16 U.S.C. 1), is directed to conserve the scenery, natural and historic objects and
wildlife and to provide for the enjoyment of these resources unimpaired for current and future
generations.
       The Wilderness Act of 1964 asserts wilderness areas will be administered in such a
manner as to leave them unimpaired and preserve them for the enjoyment of future generations.
       NFS Management Policies (2006) guide all NFS actions  including natural resources
management. In general, the NFS Management Policies reiterate the NFS Organic Act's
mandate to manage the resources "unimpaired."
U.S. Fish and Wildlife Service
       On endangered species, Title 16 USC Chapter 35 Section 1531 states "The Congress
finds and declares that— these species offish, wildlife , and plants are of esthetic , ecological,
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educational, historical, recreational, and scientific value to the Nation and its people and that all
Federal departments and agencies will use their authorities to conserve threatened and
endangered species.
       The United States Fish and Wildlife Service (FWS) manages the National Wildlife
Refuge System lands to "...ensure that the biological integrity, diversity, and environmental
health of the Systems are maintained for the benefit of present and future generations of
Americans." 16U.S.C. Section 668dd(a)(4)(B)(1997).
U.S. Forest Service
       The National Forest units are managed consistent with Land and Resource Management
Plans (LRMPs) under the provisions of the National Forest Management Act (NFMA). 16
§U.S.C. 1604 (1997).  LRMPs are, in part, specifically based on recognition that the National
Forests are ecosystems and their management for goods and  services requires an awareness and
consideration of the interrelationships among plants, animals, soil, water, air, and other
environmental factors within such ecosystems. 36 C.F.R. §219.1(b)(3)
       Any measures addressing Air Quality Related Values (AQRV) on National Forest
System lands will be implemented through, and be consistent with, the provisions of an
applicable LRMP or its revision (16 U.S.C. §1604(i)). Additionally, the Secretary of Agriculture
must prepare a Renewable Resource Program that recognizes the need to protect and, if
necessary, improve the quality of air resources. 16 U.S.C. §1602(5)(C).
       AQRVs in Wilderness areas may receive further protection by the previously mentioned
1964 Wilderness Act. For Wilderness Areas in the National Forest System, the Act's
implementing regulations are found at 36 C.F.R. §293 requiring these Wilderness Areas be
administered to preserve and protect [their] wilderness character.
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United Nations Economic Commission for Europe (UNECE)
       In many European countries a critical loads framework is used to determine a level of
damages to ecosystem services from pollution that is legally allowed.  The term critical load is
used to describe the threshold of air pollution deposition that causes a specified level of harm to
sensitive resources in an ecosystem. A critical load is technically defined as "the quantitative
estimate of an exposure to one or more pollutants below which significant harmful effects on
specified sensitive elements of the environment are not expected to occur according to present
knowledge" (Nilsson and Grennfelt,  1988). The determination of when a harmful effect
becomes "significant" may be in the view of a researcher or through a policy development
process.  Critical loads have been modeled by individual countries and submitted to the UNECE
(in cases where countries have not submitted their own critical loads those  loads have been
calculated for them) and are being used to support international emissions reduction agreements
including the 1999 Gothenburg protocol and the National Emission Ceiling Directive  of the
European Commission.  Figure 3-2 shows critical loads for eutrophication (nitrogen) and
acidification (sulfur) that protect 95% of forests, seminatural vegetation or surface waters in
Europe.
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     CLnutfN} (5th percentile }
AII ecosystem s    CLmax(S) (5th percentile )
      eqha'V
      • <200
      • 200-400
      D400-700
      D700- 1000
      DlOOO- 1500
      D"> 1500
All ecosystem
Figure 4-2   European maps of eutrophication (left) and acidification (right) which protect
95% of natural areas in 50x50 km2 European Monitoring and Evaluation Programme grid.
[Red shaded areas illustrate grid cells where deposition needs to be lower than 200 eq ha"1 a"1 to reach protection
targets Specifically protection from eutrophication or an ANC lower than 20ueq/L for acidification.Source: Critical
Load, Dynamic Modelling and Impact Assessment in Europe CCE Status Report 2008 available at
http://www.pbl.nl/en/publications/2009/Critical-load-dynamic-modelling-and-impact-assessment-in-Europe-CCE-
Status-Report-2008.htmlhttp://www.pbl.nl/en/publications/2009/Critical-load-dynamic-modelling-and-impact-
assessment-in-Europe-CCE-Status-Report-2008.html]
       The Coordination Center for Effects, a working center for the Working Group on Effects
of the Convention on Long Range Transboundary Pollution, in the 2008 status report shows
calculated critical loads based on an ANC target of 20|ieq/L for the protection of brown trout.
Individual countries have also set ANC targets for other species for example Norway uses a
critical load of 30|ieq/L for Atlantic salmon (Jenkins et al, 2003) .

4.2    ECOSYSTEM SERVICES AND ADVERSITY TO PUBLIC WELFARE
       An additional concept that may be useful in considering the issue of adversity to public
welfare is ecosystem services. In the next section the concept of ecosystem services, its
relationship to adversity and public welfare  within the context of this review are explained.
       Ecosystem services can be generally defined as the benefits individuals and organizations
obtain from ecosystems. Ecosystem services can be classified as provisioning (food and water),
regulating (control of climate and disease), cultural (recreational, existence, spiritual,
educational), and supporting (nutrient cycling) (MEA 2005). Conceptually, changes in
ecosystem services may be used to aid in characterizing a known or anticipated adverse effect to
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public welfare. In the context of this review, ecosystem services may also aid in assessing the
magnitude and significance to the public of a resource and in assessing how oxides of nitrogen
and sulfur concentrations and deposition may impact that resource. The relationship between
ecosystem services and public welfare effects is illustrated in Table 4-1.
            Table 4-1. Crosswalk between Ecosystem Services and Public Welfare Effects.
  Public Welfare Effect         Ecosystem Service            Service Category

Soils
Water
Crops
Vegetation
Wildlife
Climate
*Personal Comfort and
Wellbeing
Nutrient Cycling, Water
Filtration
Drinking water, Recreation,
Aesthetic , Nonuse
Food, Fuel Production, Forest
Products
Food, Forest Products,,
Recreation, Aesthetic,
Nonuse**
Recreation, Food, Nonuse**
Climate Regulation including
carbon sequestration,
denitrification product
emissions, effects on albedo,
biogenic emissions, and
microclimate effects

Supporting, Provisioning
Provisioning, Cultural
Provisioning
Provisioning, Cultural
Cultural, Provisioning
Regulating

       *A11 ecosystem services contribute to personal comfort and wellbeing.
       ** Nonuse values include existence, preservation, and bequest values.
       EPA has defined ecological goods and services for the purposes of a Regulatory Impact
       Analysis as the "outputs of ecological functions or processes that directly or indirectly
contribute to social welfare or have the potential to do so in the future. Some outputs may be
                                          4-11

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bought and sold, but most are not marketed" (US EPA 2006).  Additionally Executive Order
12866 requires a regulatory Impact Analysis for any rule considered "economically significant"
and defines significant as a rule having $100 million or more in impacts. Though this is not a
definition specifically for use in the NAAQS process it may be a useful one in considering the
scope of ecosystem services and the effects of air pollutants upon those services. Especially
important is the acknowledgement that it is difficult to measure and/or monetize the goods and
services supplied by ecosystems. Valuing ecological benefits, or the contributions to social
welfare derived from ecosystems, can be challenging as noted in EPA's Ecological Benefits
Assessment Strategic Plan (US EPA 2006) and the Science Advisory Board report "Valuing the
Protection of Ecological Systems and Services" (US EPA, 2009). It can be informative in
characterizing adversity to public welfare to attempt to place an economic valuation on the set of
goods and services that have  been identified with respect to a change in policy however it must
be noted that this valuation will be incomplete and illustrative only. The stepwise concept
leading to the valuation of ecosystem  services is graphically depicted in Figure 4-3.
                                          4-12

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                               Ecosystems
                         Ecological goods and services
                             affected by the policy
                       Planning and problem formulation
                       '
                       '
                            Goods and services
                                identified
   Ecological analysis
Goods and services
    quantified
                             Economic analysis
                                   Goods and
                                  services not
                                   identified
                           Goods and
                            services
                           monetized
 Identified
 goods and
services not
 quantified
                            Quantified
                            goods and
                            services not
                            monetized
Figure 4-3. Representation of the benefits assessment process indicating where some ecological
benefits may remain unrecognized, unquantified, or unmonetized. (Source: EBASP USEPA
2006).

       A conceptual model integrating the role of ecosystem services in characterizing known or
anticipated adverse effects to public welfare is shown in Figure 4-4. Under Section 109 of the
CAA, the secondary standard is to specify a level of air quality that is requisite to protect public
welfare. For this review, the relevant air quality indicator is interpreted as ambient oxides of
nitrogen and sulfur concentrations that can be linked to levels of deposition for which there are
ecological effects that are adverse to public welfare. The case study analyses (described in
Chapters 4 and 5 of the REA) link deposition in sensitive ecosystems (e.g., the exposure
pathway) to changes in a given ecological indicator (e.g., for aquatic acidification, changes in
acid neutralizing capacity [ANC]) and then to changes in ecosystems and the services they
provide (e.g., fish species richness and its influence on recreational fishing). To the extent
possible for each targeted effect area, ambient concentrations of nitrogen and sulfur (i.e.,  ambient
air quality indicators) were linked to deposition in sensitive ecosystems (i.e., exposure
                                            4-13

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pathways), and then deposition was linked to system response as measured by a given ecological
indicator (e.g., lake and stream acidification as measured by ANC). The ecological effect (e.g.,
changes in fish species richness) was then, where possible, associated with changes in ecosystem
services and their public welfare effects (e.g., recreational fishing). We recognize that there is a
certain amount of natural change in ecosystems over time that can affect the level of acidity and
the response of the ecosystem to additional acid and nutrient inputs.  However, this review is
focused on the impact of anthropogenic nitrogen and sulfur deposition given the existing state of
non-anthropogenically determined ecosystem characteristics, and as such we essentially hold
these other factors as "fixed" for the purposes of the review.
       Knowledge about the relationships linking ambient concentrations and ecosystem
services can be used to inform a policy judgment on a known or anticipated adverse public
welfare effect. The conceptual model outlined for aquatic acidification in Figure 4-4 can be
modified for any targeted effect area where sufficient data and models are available.  For
example, a change in an ecosystem structure and process, such as foliar injury, would be
classified as an ecological effect, with the associated changes in ecosystem services,  such as
primary productivity, food availability, forest products, and aesthetics (e.g., scenic viewing),
classified as public welfare effects. Additionally, changes in biodiversity would be classified as
an ecological effect, and the associated changes in ecosystem services—productivity, existence
(nonuse) value, recreational viewing and aesthetics—would be classified as public welfare
effects. This information can then be used by the Administrator to determine whether or not the
changes described are adverse to public welfare. In subsequent sections these concepts are
applied to characterize the ecosystem services potentially affected by nitrogen and/or sulfur for
each of the effect areas assessed in the RE A.
                                           4-14

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           Ambient Air Quality                 NO X/SO x
                Indicator                   Concentrations

                                        Atmospheric N & S
           Exposure Pathway                  Deposition
                                                T
           Affected Ecosystem                  Aquatic
          Ecological Response                Acidification
          (ecological indicator )            (lake/streamANC
               VC  )  \
                                       Change in Ecosystem
            Ecological Effect            Structure & Processes
                                       (fish species richness )
\
           Ecological Benefit'             c    C^ang|in .
             Welfare Effect              Ecosystem Services
                                        (recreational fishing )
Figure 4-4.   Conceptual model showing the relationships among ambient air quality indicators
and exposure pathways and the resulting impacts on ecosystems, ecological responses, effects
and benefits to characterize known or anticipated adverse effects to public welfare.
       These concepts can also be applied to the programs described in section 4.1. National
parks represent areas of nationally recognized ecological and public welfare significance, which
are afforded a higher level of protection. Therefore, staff has also focused on air quality and
deposition in the subset of national park sites and important natural areas. The spatial
relationships between sensitive regions, Class 1 areas, federal and state public lands, and
nitrogen deposition levels are illustrated in Figures 4-5 and 4-6. Please note that the scale of
deposition levels is different for the two maps to allow greater differentiation of the deposition in
                                            4-15

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the western US.
       Weight in grams of elemental nitrogen and sulfur were
       extracted from the various species reported by both
       CMAQ (dry) and NADP (wet) by dividing by the species'
       molecular weight. These values were then converted
       to equivalents of either sulfur or nitrogen and combined.
       The source of the federal and state public recreation
       areas is the Conservation Biology Institute. The lands
       were filtered to remove non-recreation areas as well
        as those less than 64,000 acres
             Combined N and S

             eq/ha/yr

             ^B 1,219-2,652

                | 1,054- 1,218

             |     | 926-1,053

             |     | 835 - 925

             |     | 753 - 834

                ^\ 607 - 752

             ^H 322 - 606
250
500
Figure 4-5.    Locations of Eastern U.S. Public Lands relative to deposition of nitrogen and
sulfur in sensitive aquatic areas.  Source 2005 CMAQ and NADP. Note: Total N includes reduced nitrogen
forms.
                                                     4-16

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          Combined N and S
          ^H  1,000-1,337
          ^B  600 - 999
            ^|  300 - 599
          |    |  250 - 299
          |    |  220 - 249
          |    |  195-219
            ^\  175-194
          ^B  140-174
          ^•69-139
              250
                        500
\Afeight in grams of elemental nitrogen and sulfur were
extracted from the various species reported by both
CMAQ (dry) and NADP (wet) by dividing by the species
molecular weight. These values were then converted
to equivalents of either sulfur or nitrogen and combined.
The source of the federal and state public recreation
areas is trie Conservation Biology Institute.  The lands
were filtered to remove non-recreation areas as well
as those less than 64,000 acres.
Figure 4-6.    Location of Western U.S. Public Lands relative to deposition of nitrogen and
sulfur. Source 2005 CMAQ and NADP.  Note: Total N includes reduced nitrogen forms.
                                                       4-17

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4.3    APPLYING ECONOMIC VALUATION TO ECOSYSTEM SERVICES
       As discussed earlier in this document, a secondary NAAQS is required to be set at the
"level(s) of air quality necessary to protect the public welfare from any known or anticipated
adverse effects".  As part of the effort to determine the standard, EPA linked the changes in the
ambient air concentrations of oxides of nitrogen and sulfur to the changes in ecosystem services
and ultimately to changes in public welfare (U.S. EPA, 2009). The difficulty in the monetization
for ecosystem services has been previously emphasized.  This difficulty necessitates focusing on
a subset of services for economic valuation.  And although economics on its own cannot
determine what level of impact on public welfare is "adverse," economics can be helpful in the
context of a secondary NAAQS for determining the degree to which improvements are beneficial
to public welfare and illustrating and aggregating those impacts.l
       An ecosystem service framework provides a structure to measure changes in public
welfare from changes in ecosystem functions affected by air pollution. EPA's Risk Assessment
for this rulemaking defines ecosystem services as "the ecological processes  or functions having
monetary or nonmonetary value to individuals or society at large" (EPA 2009). The discipline of
economics provides a useful approach for summarizing how the public values changes in the
services provided by the environment.  An ecosystem services framework (with or without
valuation) can characterize and describe how changes in ecosystem function affect public
welfare and provide measures of changes in public welfare where those affects  can be quantified.

4.3.1   Economics as a Framework to Illustrate Changes in Public Welfare
       Economics can provide a framework to illustrate how public welfare changes in response
to changes in environmental quality by quantitatively linking changes in ecosystem services to
preferences. Economics assumes that the choices that individuals make reflect their preferences
over certain outcomes and that, generally speaking, they will make choices that, in expectation,
will make them as well off as possible given their resources. An individual's preferred outcomes
may include not just their own use and enjoyment of an ecosystem but also preservation and
1 Section 109 of the Clean Air Act forbids consideration of the compliance costs of reducing pollution when setting
a NAAQS. However, there is no prohibition regarding the consideration of the monetized impacts of welfare effects
occurring due to levels of pollution above alternative standards in evaluating the adversity of the impacts to public
welfare.  Ecosystem services can be characterized as a method of monetizing the impacts of the air pollution.
Although a separate regulatory document quantifying the costs and benefits of attaining a NAAQS is prepared
simultaneously, this document is not considered when selecting a standard.
                                           4-18

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bequest value. In economics, revealed and stated preference methods are used to observe the
choices individuals make to understand the outcomes individuals prefer. What individuals are
willing to give up for an outcome is their willingness-to-pay (WTP) for that outcome. An
example of an outcome is an improvement in an ecosystem service. Often, to provide
comparability to other goods and services, in economics these tradeoffs are framed relative to
dollars for convenience.
       Economics could inform the Administrator by valuing and characterizing the changes in
public welfare from changes in the quantity and quality of ecosystem services.  Overall, this
assessment intends to characterize changes in ecosystem services from a scientific perspective
using effects on ecosystem structures and functions or ecosystem integrity. Economics then
estimates the  effect on public welfare of these changes in the quantity and quality of ecosystem
services using willingness to pay as a measure of this effect. For example, a decrease in a
particular bird species can be characterized by its effect on the ecosystem's structure and
function, while from an economic perspective, the  effects would be based on the impact on
public welfare or the value the public places on that species. A simple example is a comparison
between a decrease in a bird species that  is relatively unknown compared to a decrease in a very
prominent species (e.g. bald eagle). The  public is likely to have a higher WTP to avoid the latter,
and thus the decrease would affect the public welfare more, even if the changes in the two bird
species generally have the same impact on an ecosystem's structure or function.
       There are important complications with using preferences to understand the effect of
pollution on public welfare. For example, while the field of economics generally assumes that
public preferences are the paramount  consideration; care must be taken that these preferences
may change when the public receives new information. Evaluation of public preferences should
take place under conditions of full information. If individuals do not understand how pollution
will affect ecosystem services, or even how those ecosystem services affect their quality of life,
then they will have a difficult time valuing changes in those services. Similarly, it may be very
difficult and time-consuming for individuals to learn and understand how changes in particular
ecosystem services may affect them, in part because typically there are  significant
interdependences within an ecosystem. Because of this complexity, individuals may implicitly
value a species, or habitat, or ecosystem function because it supports an ecosystem service that
they do clearly value.  Furthermore, the public also has limited understanding regarding
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irreversibilities, tipping points, and other more complex aspects of ecosystems, which limits the
ability to adequately value these ecosystems.2 In addition, where and when a change in an
ecosystem takes places is crucial for characterizing the associated change in an ecosystem
service, and will also affect the value the public places on that change.
       The fact that collective action activities are being undertaken by communities, Non-
Governmental Organizations (NGO's) and States underscores the fact that there is a societal
demand for further improvement to the quality of many water bodies which have been impaired
by acidic deposition.3 However, as illustrated below they provide insufficient quantitative
evidence as to what society willingness to pay to reduce lake and stream acidity because it is
difficult to separately identify individual preferences from the actions of the group.

4.3.2  The Role of Economics in Defining Adversity
       If economic valuation can establish a significant loss to public welfare, then this can
provide strong  support for a determination of an "adverse" effect. However, there is neither an
economic definition of how much  loss in public welfare is adverse nor an economic definition of
adversity. While an economist might consider a particular scenario adverse because it might
imply some harm or potential for improvement, there is no specific threshold level when a loss in
welfare (e.g. loss in dollars) becomes adverse. An individual might be willing to give up some
of their resources to  avoid a threat or negative outcome (i.e., willing to pay to avoid a particular
outcome). According to economic theory, if an individual is willing to give up something to
avoid the outcome, then imposing  the outcome on the individual must make them worse off, at
which point an economist might describe the outcome as adverse. However, the amount an
individual is willing to pay to avoid the outcome may or may not rise to the level of harm that the
Administrator interprets as "adverse" to public welfare. At the same time, an economic
valuation that shows that there are substantial damages from current levels of acidification or
nutrient enrichment would provide strong evidence for finding that current impacts are adverse
2 While the public may not fully appreciate the interdependencies within ecosystems, they can learn them, but again
it may be costly to do so. It is possible for individuals to value outcomes that are irreversible or result in discrete
changes (i.e., tipping points) in the quality and quantity of ecosystem services. Avoiding irreversible outcomes
should be and are more valued by individuals than outcomes that are not irreversible (Arrow and Fischer, 1974).
 However, one must recognize that often times reducing acidity is often part of a larger effort to generally improve
the quality of a water body. Therefore, separating out the portion of people' s desire to just to reduce acidity from
the more general improvements is difficult.
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to public welfare. In summary, economic analysis (particularly valuation) can provide useful
information for the Administrator as to the interpretation of the word "adverse" in the context of
public welfare, but it does not provide a complete set of information needed to make that
determination.

4.3.3  Collective Action as an Indicator of Adversity
       Typically, economics uses information on willingness to pay for improved environmental
quality that is gathered from observing individuals' market behavior (revealed preference) or that
they provide through surveys (i.e., stated preference methods). The analyses in the following
sections use revealed and stated-preference information to quantify a portion of the social costs
of current levels of acidification and nutrient enrichment. However, the studies supporting these
analyses evoke specific contexts and thus the findings may not be generalizable across all of
those affected by acidification or nutrient enrichment.4 An alternative source of revealed
information on the damages caused by acidification can be found in the behavior of groups.
Often groups collectively make choices to engage in activities that improve the collective
welfare of the group. For example, a community around an acidified lake might undertake
activities designed to improve the quality of that lake, including purchasing lime, to use as a tool
to reduce the acidity of the lake. These collective decisions can be used to gain insights into how
people value improvements to ecosystem services. However, there are many obstacles to
collective actions,  including problems of organization, free ridership and others (Olson 1965)
that make it difficult to use the actions of organizations to interpret individuals' preferences.
       In addition to communities,  states may also take actions to increase the quality of their
impaired lakes. Non-Governmental Organizations or  advocacy groups, as well may organize
support for,  and/or directly undertake, activities to improve lake and stream quality on behalf of
its members/donors. How individual's preferences are expressed through these collective actions
is discussed below. For brevity, this discussion will focus on collective efforts to reduce acidity
of lakes and streams by Communities, Nongovernmental Organizations and States.
4 Even in the case where the existing studies provide a reliable characterization of the effects of acidification or
nutrient enrichment on a limited number of individuals, it is advisable to make use of corroborating data and studies
when such information is available.
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Communities
       In cases where property rights to a resource are well defined, collective action is more
likely to take place, as individuals have greater ownership and control over the affected resource.
Rights to use a lake, as well as, mandatory membership in a lake association is often written right
into the deed of properties which abut or surround a lake, giving these property owners more
control over the resource. This mechanism of granting rights and responsibilities to the lake
encourages better management of the lake resource by remedying unrestricted access and free
rider problems.  This coupling of the costs of resource improvements with their benefits
encourages individuals to maintain the quality of the resource.
       There have been several documented instances where communities (particularly
Homeowners Associations) have spent time and money to improve the quality of a lake. These
include actions to combat acidity, eutrophication, invasive species (e.g. Zebra Mussels) and other
problems. The Lake Wononscopomuc Association in Salisbury, Connecticut is a typical
example (Mayland 2009.)  They spend their own funds to hire scientific consultants to survey
and test the lake water (for e.g. pH., dissolved oxygen, visibility, and many other factors related
to the lake's condition) and recommend  management strategies to improve the quality of the
lake.  Likewise, in Georgia, the Berkeley Lake Homeowners Association (BLHA) is a non-profit
homeowner association dedicated to protecting Berkeley Lake.  BLHA is typical of many other
home owners associations with access to a lake, in that they are also concerned with and
managing acidity, eutrophication, invasive species and a whole host of more mundane upkeep
and maintenance issues (Hunkapiller 2006.)  BLHA recognizes the relationship between lake
acidity and resident's enjoyment of the lake's fishing swimming and aesthetics.
Nongovernmental Organizations
       Nongovernmental Organizations (NGOs) or Advocacy Groups organize individuals and
smaller groups thereby reducing the transaction costs associated with individual's desires to
advance a specific goal.  For example, Living Lakes, Inc. (LLI) is a not-for profit organization
which sponsors an applied aquatic resources restoration demonstration program for acidified
waters. In the late 1980's LLI began evaluation of seven different liming technologies on 22
lakes  and 10 streams in 6 states. Lakes and streams have been treated in the states of
Massachusetts, Rhode Island, New York, Pennsylvania, Maryland and West Virginia (Brocksen
and Emler 1988.) Likewise, sportsman groups such as Trout Unlimited, as well as,  smaller local
                                          4-22

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groups, have an interest in improving or maintaining the quality of lakes and streams.  Trout
Unlimited is well known for these activities and is discussed further later on.  However, several
smaller, localized groups also work to decrease aquatic acidification. One of these is the
Mosquito Creek Sportsman's Association in Pennsylvania. Mosquito Creek and its main
tributary Gifford Run were once famous for naturally reproducing wild brook and brown trout.
However, since the early 1960's, the pH of the stream steadily declined due to acid rain. As a
result, wild brook trout and brown trout have substantially declined in the watershed (Hoover
and Rightnour, 2002.) Aerial liming was undertaken as part of an overall watershed restoration
program that included constructed wetlands, forest liming, and in-stream liming to improve this
fishery and provide increased opportunities for public recreation in the region. Fifty tons of lime
were applied in the headwaters of Mosquito Creek Watershed. This liming project was part of
the Mosquito Creek Sportsman's Association's efforts to improve the water quality  of the 90
square mile watershed located in Clearfield and Elk counties. However, while the project first
phase and the other ongoing phases of the overall restoration project have been initiated by the
Mosquito Creek Sportsman Association, they received technical  support from multiple public,
private and other non-profit groups.5  "A benefit/cost analysis was prepared on the  four
implementation phases of this project. Costs were based on alkaline deficiencies  and the
additional costs determined for the technologies. Benefits were estimated as returns on direct
recreational use losses and community willingness-to-pay. Full restoration of the watershed is
estimated to cost approximately $3.4 million over 15 years, for an annualized cost of $229,000,
or $5,400 per mile per year for 42 miles of potential improvements. Expected returns range from
$1.2 million per year for recreational use to $6.1 million per year for total community
willingness-to-pay. It was concluded that restoration is technically feasible and economically
beneficial for the Mosquito Creek watershed, and it is recommended that planned projects and
the remainder of the progressive restoration plan be implemented." (Hoover and Rightnour,
2002)
 These included: the Pennsylvania State University Environmental Resource Research Institute, Pennsylvania
Game Commission, Pennsylvania DCNR Bureau of Forestry, Pennsylvania Fish and Boat Commission,
Pennsylvania Department of Environmental Protection, Pennsylvania Department of Corrections Quehanna Boot
Camp, Wood Duck Chapter Trout Unlimited, Canaan Valley Institute, Clearfield County Conservation District and
Pennsylvania USD A Natural Resource Conservation Service.
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States
       Several states including Vermont, New Hampshire, New York and Tennessee have
developed Total Maximum Daily Loads (TMDLs) for lakes impaired for acidification in their
jurisdictions. As mentioned in the previous section regarding the TMDL for the Chesapeake Bay
the applicable water quality standards and designated uses are set by the states with public
participation.  Although most states set their standard either by legislation or regulations, in at
least one case, specifically New York, the designated uses and water quality standard are part of
the state constitution.  The Adirondack Forest Preserve is required to be "forever kept as wild
forest lands."  New York has interpreted this to mean that the waters included in the preserve are
required to be kept in natural conditions. To this end New York has chosen to set a tiered TMDL
that allows interim water quality targets in order to address the reality that some lakes in the
Adirondacks will naturally have a pH that does not meet the state's water quality standards. For
lakes that can meet the standards the state has chosen to set the water quality standard for pH
values above 6.5. New Hampshire has chosen to set their water quality target at an ANC of
60|ieq/L that, according to the TMDL document, corresponds to a pH of 6.5. Vermont, in a
similar process has chosen a target  ANC of 50|ieq/L. In Tennessee the state faces a similar
problem as New York in trying to set levels to protect streams Great Smoky Mountains National
Park which include some naturally  acidic streams.  Accordingly they have set site specific ANC
targets where data is available to do so and chosen to target an ANC of 50|ieq/L as a default
value where data is not available. The Tennessee TMDL is a partnership between the state and
the National Park Service which is  sharing the data collection and modeling activities with
academic institutions.
       In each case the implementation sections of these  TMDLs cites the fact that the sources
of pollution responsible for the degradation of water quality in the named lakes and streams are
not located within the jurisdiction of the state. Each state has called on EPA to require regional
or national decreases in acidifying deposition. Vermont goes so far as to say "In short,
implementation of this TMDL is primarily the responsibility of EPA.... This TMDL sets out
clear  endpoints to guide EPA actions.  However, in the absence of vigorous efforts by EPA to
bring about reductions in acid emissions from out-of-state sources this TMDL will merely have
been a paper exercise."
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Public-Private Collaborations
       In some cases, industry, government and private efforts partner to reduce the acidity of a
lake or stream.  In one such instance in 2005, The U.S. Forest Service used helicopters to apply
200 tons of limestone sand into the St. Marys River and its tributaries to lower acid levels in one
of Virginia's prime trout fisheries to mitigate the impacts of acidification until a long-term
solution to acidification is found. The NGO, Trout Unlimited was one of the partners in the
liming project, while Dominion Virginia Power provided $10,000 for the liming project
(Associated Press 2005.) In another partnership, Living Lakes participated in a project in the
Woods Lake Watershed  in the Adirondack region of the state of New York that was co-
sponsored by the Electric Power Research Institute (EPRI), which sponsored the original
research on lake liming in the Adirondacks (Scheffe et al,  1986). There are many such
examples, where all three of these types of groups partner on the same project.

4.4    EFFECTS OF ACIDIFICATION AND  NUTRIENT ENRICHMENT ON
ECOSYSTEM SERVICES
       The process used to link ecological indicators to ecosystem services is discussed
extensively in Appendix 8 of the REA.  In brief, for each case study area assessed the ecological
indicators were linked to an ecological response that was subsequently linked, to the extent
possible, to associated services.  For example in the case study for aquatic acidification the
chosen ecological indicator is ANC which can be  linked to the ecosystem service of recreational
fishing as illustrated in the conceptual model shown in Figure 4-7. Although recreational fishing
losses are the only service effects that can be independently quantified or monetized at this time,
there are numerous other ecosystem services that may be related to the ecological effects of
acidification.
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     Acidifying Inputs
      ,S Depositor,
                         Ecological
                                        /mpacts Qfi Ecosystem Endsxiints
                                                                     Affected Ecosystem Services
     Acidification of Soil
       Leachate to
      Surface Water




Low p
At
i
cation:
Hand
C
Mobilization of
Alym num:
Decl

	 1 Declines in
"""*] Aquatic Biola
Fitness:
Reduced
grovMi
development,
and
repfoduction
t

Rec
S«
^ A bun
Divers
Rcl

nes in
c Biota:
uced
vies
dance.
ity, and
7/5655

»• Decl nes in
Terrestnai
Nearsnore
Biota
                                                                      Prov..s'oning Services
                                                                      •procuct or* for commercial
                                                                      and subsistence fishing
Cultural Services
•recreational fishmg
•waterfowl hunting
•aesthetic enjoyment
•no nuse services
                                                                      Regulating services
                                                                      •biological control
Figure 4-7.    Conceptual model linking ecological indicator (ANC) to affected ecosystem
services.  The red arrows highlight the path to monetization of recreational fishing effects.
Nonuse services include biodiversity,  habitat preservation, existence, and bequest value.

        While aquatic acidification is the focus of this policy assessment, the other effect areas
analyzed in the REA still merit some discussion in view of the fact that these ecosystems are
being harmed by nitrogen and sulfur deposition and will obtain some measure of protection with
any decrease in that deposition regardless of the reason for the decrease. In next four sections we
summarize the current levels of specific ecosystem services for aquatic and terrestrial
acidification, and aquatic and terrestrial nutrient enrichment. We also present results of analyses
that have attempted to quantify and monetize the harms to public welfare, as represented by
ecosystem services, due to nitrogen and sulfur deposition.

4.4.1   Aquatic Acidification
        Acidification primarily affects the ecosystem services that are derived from the fish and
other aquatic life found in these surface waters (REA, Section 5.2.1.3). In the northeastern
United States, the surface waters affected by acidification are not a major source of
commercially raised or caught fish; however, they are a source of food for some recreational and
subsistence fishers and for other consumers. Although data and models are available for
examining the effects on recreational fishing, relatively little data are available for measuring the
effects on subsistence and other consumers.  For example, although there is evidence that  certain
                                             4-26

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population subgroups in the northeastern United States, such as the Hmong and Chippewa ethnic
groups, have particularly high rates of self-caught fish consumption (Hutchison and Kraft, 1994;
Peterson et al, 1994), it is not known if and how their consumption patterns are affected by the
reductions in available fish populations caused by surface water acidification.
       Inland surface waters support several cultural services, such as recreational fishing,
aesthetic and educational services; however, Banzhaf et al (2006) has shown that non-use
services, which include existence (protection and preservation with no expectation of direct use)
and bequest values, are arguably a significant source of benefits from reduced acidification. The
areas of the country containing the most sensitive  lakes and streams are New England, the
Adirondack Mountains which are part of the Adirondack Forest Preserve - that have been set
aside to be kept "forever wild" see PA Sec 4.3.4, the Appalachian Mountains (northern
Appalachian Plateau  and Ridge/Blue Ridge region) and the Upper Midwest. Within the
Adirondack Mountains approximately 8% of the lakes were considered acidic and in the northern
Appalachian Plateau  and Ridge/Blue Ridge 6 - 8% of the streams (ISA 3.2.4.2 and REA 4.2.2).
Recreational fishing in lakes and streams is among the most popular outdoor recreational
activities in the northeastern United States. Data from the 2006 National Survey of Fishing,
Hunting, and Wildlife Associated Recreation (FHWAR) indicate that more than 9% of adults in
this part of the country participate annually in freshwater fishing with 140 million freshwater
fishing days. Based on studies conducted in the northeastern United States, Kaval and Loomis
(2003) estimated average consumer surplus values per day of $35 for recreational fishing (in
2007 dollars). Therefore, the implied total annual value of freshwater fishing in the northeastern
United States was $5 billion in 2006. We recognize that embedded in these numbers is a degree
of harm to recreational fishing services due to acidification that has occurred over time.  These
harms have not been  quantified on a regional scale. However given the magnitude of the
resource, the length of time nitrogen and  sulfur have been affecting freshwaters in the northeast
and the level of monetary damages calculated for the case study in the Adirondacks described in
the next section we would expect these damages to be significant.
       In general, inland surface waters such as lakes, rivers, and streams provide a number of
regulating services, playing a role in hydrological  regimes and climate regulation. There is little
evidence that acidification of freshwaters in the northeastern United States has significantly
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degraded these specific services; however, freshwater ecosystems also provide biological control
services by providing environments that sustain delicate aquatic food chains.
       The toxic effects of acidification on fish and other aquatic life impair these services by
disrupting the trophic structure of surface waters (Driscoll et al, 2001). Although it is difficult
to quantify these services and how they are affected by acidification, it is worth noting that some
of these services may be captured through measures of provisioning and cultural services. For
example, these biological control services may serve as "intermediate" inputs that support the
production of "final" recreational fishing and other cultural services.

4.4.2  Value of Current Ecosystem Service Impairments Due to Aquatic Acidification
       In the previous section we described the ecosystem services that are most likely to be
affected by N and S deposition and summarized evidence regarding the current magnitude and
values of recreational fishing services, the degree to which these  services are impaired by
existing NOX/SOX levels has not been quantified. To address this  limitation, the REA (Appendix
8) provides insights into the  magnitude of ecosystem service impairments.  The REA provides
quantitative estimates of selected ecosystem services impairments or enhancements for three
main categories of ecosystem effects - aquatic acidification, terrestrial acidification, and aquatic
nutrient enrichment6. Within these three categories, the  selection of specific ecosystem services
for more in-depth analysis depended primarily on the expected magnitude of impairments and on
the availability of appropriate data and modeling tools.
       The analysis of ecosystem service impairments due to aquatic acidification builds on the
case study analysis of lakes in the New York Adirondacks. In this study estimates of changes in
recreational fishing services are determined, as well as changes more broadly in "cultural"
ecosystem services (including recreational, aesthetic, and nonuse services). First, the MAGIC
model (REA, Appendix 8, Sec 2.2)  was applied to 44 lakes to predict what ANC levels would
be under both "business as usual" conditions (i.e., allowing for some decline in deposition due to
existing regulations) and pre-emission (i.e., background) conditions. These model runs assumed
a 2010 "zero-out" emissions scenario (where all N and S deposition is eliminated) with a
projected lag time between the elimination of emissions  to observed improvement in ANC of 10
years thus benefits results were calculated for the year 2020.  These predictions were then
6 Estimates for terrestrial nutrient enrichments were not generated due to the limited availability of necessary data
and models for this effect category.

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extrapolated to the full universe of Adirondack lakes.  Table 4-2 reports the number of
"impacted" lakes in each year, where impact means that the lake is predicted to be below the
ANC threshold under business-as-usual and above the threshold under pristine conditions.

       Table 4-2. Count of Impacted Lakes.
ANC Threshold
(in jieq/L)
20
20
20
50
50
50
100
100
100
Year
2020
2050
2100
2020
2050
2100
2020
2050
2100
Lake Count
107
95
74
244
222
200
430
404
354
       Note: There are 1,076 lakes in the "Adirondack Region".
       Second, to estimate the recreational fishing impacts of aquatic acidification in these lakes,
an existing model of recreational fishing demand and site choice was applied.  This model
predicts how recreational fishing patterns in the Adirondacks would differ and how much higher
the average annual value of recreational fishing services would be for New York residents if lake
ANC levels corresponded to background (rather than business as usual) conditions.  Table 4-3
summarizes the results and the present value of benefits and annualized benefits at 3 and 7%
discount rates.
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       Table 4-3.  Present Value and Annualized Benefits of Recreational Fishing to NY
       Residents, Adirondack Region.
ANC
Threshold
(in (ieq/L)
20
50
100
Present Value Benefits3
(in million of 2007 dollars)
3% Discount
Rate
$142.59
$285.15
$298.67
7% Discount
Rate
$60.05
$114.18
$120.61
Annualized Benefitsb
(in million of 2007 dollars)
3% Discount
Rate
$4.46
$8.91
$9.33
7% Discount
Rate
$3.94
$7.49
$7.91
a Annual benefits for 2010 to 2100 discounted to 2010.
b Present value benefits annualized over 2009-2100.
        Current annual impairments are most likely of a similar magnitude because, although
current NOX/SOX levels are somewhat higher than those expected in 2020 (under business as
usual - given expected emissions controls associated with Title IV regulations but no additional
nitrogen or sulfur controls), and the affected NY population is also somewhat smaller (based on
U.S. Census Bureau projections).
       To estimate impacts on a broader category of cultural (and some provisioning)
ecosystem services, results from the Banzhaf et al (2006) valuation survey of New York
residents were adapted and applied to this  context. The survey used a contingent valuation
approach to estimate the average annual household WTP for future reductions (20% and 45%) in
the percent of Adirondack lakes impaired by acidification.  The focus of the survey was on
impacts on aquatic resources. Pretesting of the survey indicated that respondents nonetheless
tended to assume that benefits would occur in the  condition of birds and forests as well as in
recreational fishing. The survey that measured the benefits of 20% of the lakes improving
indicated that terrestrial benefits were minor and econometric controls were used to adjust the
willingness to pay estimate for those that suspected that terrestrial improvements were greater
than described in the survey. The survey that measured the benefits of improving 45% of the
total number of lakes also indicated that the  benefits to forests and birds were significant.
       The WTP estimates from the two versions of the survey were then (1) scaled to reflect
predicted changes between business-as-usual and background conditions in 2020 (MAGIC lake
modeling results indicate that impaired lakes would decrease from 22 to 31% using background
conditions with ANC increasing from 20 to  50|ieq/L), and (2) aggregated across New York
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households. The scaling entails converting the average household willingness-to-pay for the
improvements described in the Adirondacks surveys to an average household willingness-to-pay
per percentage point of the total population of lakes improved.7 The results are summarized in
Table 4-4. The range of average household willingness to pay reflects the range in willingness to
pay per percentage point of lakes improved described in the two versions of the survey.
Estimates are provided at ANC 20, 50, and 100 to reflect the range of ANC discussed throughout
the REA and this document and for consistency with the Random Utility Model analysis.
       The "base" version of the survey asserts that, in the absence of any direct policy
intervention, the condition of the  1,500 unhealthy lakes and 1,500  healthy lakes is expected to
remain unchanged over the next 10 years. However, if a liming program is undertaken, it would
improve 20% (600) of the lakes in the Park relative to their expected 2014 condition without the
program. In contrast, the "scope" version describes a gradually worsening status quo without the
liming program, in which 5% (150) of the healthy lakes are expected to gradually become
unhealthy. In other words, without the program, 55% (1,650) of the lakes would be unhealthy in
2014. With the liming program, however, only 10% of the lakes would be unhealthy in 2014, so
the program improves 45% (1,350) of the lakes relative to their expected 2014 condition without
the program.
       Although scientific evidence indicates that a liming policy  would not significantly
improve the condition of birds and forests, pretesting of the survey indicated that respondents
nonetheless tended to assume that these other benefits would occur. Therefore, to make the
scenarios more acceptable to respondents, other nonlake effects were added to the two survey
versions. In the base case, the red spruce (covering 3% of the forests' area)  and two aquatic bird
species (common loon and hooded merganser) are said to be affected. In this version, the health
of birds and forests is described as unchanged in the absence of intervention, and minor
improvements are said to result from the program. In the scope version, a broader range of
damages is associated with acid rain—two additional species of trees (sugar maple and white
ash, all together covering 10% of forest area) and two additional birds (wood thrush and tree
swallow) are said to be affected. The scope version describes a gradually worsening status quo
along with large improvements due to the program.
7 Scaling is required because neither of the surveys administered by Banzhaf et al. (2006) describe improvements
that correspond exactly to the improvement scenario modeled here.

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       Each respondent was presented with one of these (base or scope) policy scenarios and
then asked how they would vote in a referendum on the program, if it were financed by an
increase in state taxes for 10 years. To estimate the distribution of WTP, the annual tax amounts
were randomly varied across respondents. Based on a detailed analysis of the survey data,
Banzhaf et al. (2006) defined a range of best WTP estimates, which were converted from 10-year
annual payments to permanent annual payments using discount rates of 3% and 5%. For the base
version, the best estimates ranged from $48 to $107 per year per household (in 2004 dollars), and
for the scope version they ranged from $54 to $154.
       To specify values for WTPAdr, these estimates were converted to 2007 dollars using the
CPI and each of them was divided by the corresponding change in the percentage of lakes that
are unhealthy (20% for the base version and 45% for the scope version). For the base version, the
WTPAdr estimates range from $2.63 to $5.87 per percentage decrease in unhealthy lakes, and for
the scope version they range from $1.32 to $3.76.
       To estimate NNY, the Census population projection for  New York for 2010 was used,
which is 19.26 million people, and this amount was divided by the ratio of population size to the
number of households in New York (2.69) in the year 2000 (assuming that this ratio stays
constant from 2000 to 2010).
       Finally, to estimate A%IL the MAGIC model results were used,  and it was assumed that
the distribution of ANC levels for these 44 lakes is representative of all  3,000 lakes in the
Adirondacks Park.  In 2020, the reduction in the percentage of lakes that are unhealthy in the
zero-out condition compared to the reference condition is 22% for the 20 ueq/L threshold. For
the 50 ueq/L, and 100 ueq/L thresholds, it is 31% and 26%, respectively. These 3% reduction
values were used as the main estimates o£A%IL.
       To estimate aggregate benefits for the zero-out scenario using the RFF survey results, it is
important to use the results from the survey version that most  closely  match this scenario.
Although both RFF survey versions use 2004 as the "current" year instead of 2010, they both use
a 10-year horizon, which corresponds to the zero-out scenario. Although no direct matches exist,
the closest correspondence is between the zero-out scenario assuming a 50 ueq/L threshold and
the RFF scope survey. Under current and future conditions with no additional policy
interventions, the RFF scope scenario assumes a small increase in unhealthy lakes from 50% to
55%, whereas the 50 ueq/L threshold is expected to result in a small decrease from 43% to 42%.
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With the program, the RFF scope survey describes a 45% decrease in unhealthy lakes, whereas
the zero-out scenario projects a 31% decrease.
       Table 4.4 reports the aggregate benefit estimates for the zero-out scenario using the 50
ueq/L threshold. As described above, the projected long-term decrease in the percentage of
unhealthy lakes (A%IL) for this scenario is 31%. Using the range of WTPAdr values from the
RFF scope survey and the projected number of New York households in 2010 the aggregate
annual benefits of the zero-out scenario  are estimated to range from $291 million to $829
million. Table 3.4 also reports aggregate benefit estimates for the zero-out scenarios using the
20 ueq/L and 100 ueq/L thresholds for ANC. Neither of these scenarios corresponds well with
the baseline descriptions of either the base or scope version of the RFF survey. The baseline
percentage of unhealthy lakes using the  20 ueq/L threshold (22%) is much lower than in either
the survey version. In contrast, the percentage using the 100 ueq/L threshold (77%) is much
higher. Nevertheless, the future reductions in the percentage of unhealthy lakes (22% and 26%)
are closest to the reductions described in the base version of the RFF survey. Therefore, the
aggregate benefits of the zero-out scenario with these thresholds are evaluated using the range of
WTPAdr values from the RFF base survey. With the 20 ueq/L threshold, the aggregate benefits
are estimated to range from $411 million to $916 million per year. With the 100 ueq/L threshold,
the aggregate benefits are estimated to range from $492 million to $1.1 billion per year.
       Table 4-4.  Aggregate Annual Benefit Estimates of Recreational Fishing to NY Residents
       for the Zero-Out Scenario.
ANC
Threshold
20 ueq/L
50 ueq/L
100 ueq/L
Reduction in
Percentage of
Unhealthy
Lakes
22%
31%
26%
Range of Average
Household WTP
per Percentage
Reduction
*$2.63
**$1.32
*$2.63
$5.87
$3.76
$5.87
Number of NY
Households
(in millions)
7.162
7.162
7.162
Range of Annual
Benefits
(in millions of 2007$)
$410.6
$291.2
$491.6
$916.4
$829.4
$1,097.2
       *Base version WTP   **Scope version WTP
       These results suggest that the value of avoiding current impairments to ecosystem
services from Adirondack lakes are even higher than the estimate, because the estimates assume
a lag of 10 years in which no benefits accrue and because the percent of impaired lakes is slightly
higher today than expected in 2020 under business-as-usual. These results imply significant
value to the public in addition to those derived from recreational fishing services. Note that the
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results are only applicable to improvements in the Adirondacks valued by residents of New
York.  If similar benefits exist in other acid-impacted areas, benefits for the nation as a whole
could be substantial. The analysis provides results on only a subset of the impacts of acidification
on ecosystem services and suggests that the overall impact on these services is likely to be
substantial.

4.4.3  Terrestrial Acidification
       In the previous chapter of this document we discussed the effects of acidifying deposition
on terrestrial ecosystems, especially forests.  These include the observed decline and/or dieback
in red spruce and sugar maple. These species are particularly sensitive to acidifying deposition
and have ranges that overlap the areas of the U.S. where some of the highest levels of acidifying
deposition occur. Additionally these species are present in the case study areas examined in the
REA.  As a result we chose to focus on red spruce and sugar maple as the species of interest for
the analysis  of ecosystem services presented in this section.
       A similar model to Figure 4-7 can be drawn for terrestrial acidification that links Be: Al
molar ratio to reduced tree growth to decreases in timber harvest, although we have less
confidence in the significance of this linkage than we do for aquatic acidification.  There are
numerous services expected to be affected but the data and methods to adequately describe those
losses does not as yet exist. These services include effects to forest health, water quality, and
habitat, including decline in habitat for threatened and endangered species, decline in forest
aesthetics, decline in forest productivity, increases in forest soil erosion and decreases in water
retention (EPA, 2009; EPA, 2009; Krieger, 2001). Forests in the northeastern United States
provide several important and valuable provisioning services, which are reflected in the
production and sales of tree products. Sugar maples are a particularly important commercial
hardwood tree species in the United States, producing timber and maple syrup that provide
hundreds of millions of dollars in economic value annually (NASS, 2008). Red spruce  is also
used in a variety of wood products and provides up to  $100 million in economic value annually
(USFS, 2006).
       Forests  in the northeastern United States are also an important source of cultural
ecosystem services, including nonuse (existence value for threatened and endangered species),
recreational, and aesthetic services (U.S. EPA, 2009; U.S. EPA, 2009). Red spruce forests are
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home to two federally listed species, the spruce-fir moss spider and the rock gnome lichen.  The
value of these two endangered species has not been estimated.
       Although we do not have the data to link acidification damages directly to economic
values of lost recreational services in forests, these resources are valuable to the public.  For
example, most recent data from the National Survey on Recreation and the Environment (NSRE)
indicate that,  from 2004 to 2007, 31% of the U.S. adult (16 and older) population visited a
wilderness or primitive area during the previous year, and 32% engaged in day hiking (Cordell et
al, n.d.). A recent study suggests that the total annual value of off-road driving recreation was
more than $9 billion, total and value of hunting and wildlife viewing was more than $4 billion
each in the Northeastern United States in 2006 (Kaval and Loomis, 2003). Table 4-5
summarizes data from the NSRE and the Fishing, Hunting,  and Wildlife Related Activity Survey
(U.S. DOT, 2007) along with average WTP estimates from Kaval and Loomis to estimate the
total value of these services in the northeast.
       Table 4-5. Annual participation and value of outdoor (forest related) activity in the
       northeast.
Recreational
Activity
Off Road Driving
Hunting
Wildlife Viewing
Participation
Rate
16
5.5
10
Activity
Days
(in
Thousands)
366,336
83,821
122,200
Avg.
WTP Per
Activity
Day
($2007)
$25.25
$52.36
$34.46
Total
Value
(in millions)
$9,250
$4,380
$4,210
         In addition, fall color viewing is a recreational activity that is directly dependent on
forest conditions. Sugar maple trees, in particular, are known for their bright colors and are,
therefore, an essential aesthetic component of most fall color landscapes. Statistics on fall color
viewing are much less available than for the other recreational and tourism activities; however, a
few studies have documented the extent and significance of this activity. For example,  Spencer
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and Holecek (2007) found that approximately 30% of residents in the Great Lakes area reported
at least one trip in the previous year involving fall color viewing. In a separate study conducted
in Vermont, Brown (2002) reported that more than 22% of households visiting Vermont in 2001
made the trip primarily for the purpose of viewing fall colors (Table 4-6).
       Table 4-6.  Summary of Studies of Fall Color Viewing .
Ecosystem Service
Fall Color Viewing

% Population
30%
22%
Population surveyed
Great Lakes area residents
Vermont visitors
Study Cited
Spencer (2007)
Brown (2002)
       Two studies have estimated values for protecting high-elevation spruce forests in the
Southern Appalachians.  These forests occur mostly in the Great Smoky Mountains National
Park, the North Carolina Park System, the Jefferson and Pisgah National Forests, and the Blue
Ridge parkway.  Kramer et al. (2003) conducted a contingent valuation study estimating
households' WTP for programs to protect remaining high-elevation spruce forests from damages
associated with air pollution and insect infestation (Haefele et al., 1991; Holmes and Kramer,
1995). While it is not possible to separate the relative damage attributable only to air pollution it
should be noted that the  insect infestation referred to in these studies is specifically damage
caused by the balsam wooly adelgid to Frasier fir trees that are part of the spruce-fir ecosystem.
       The survey presented respondents, who lived within 500 miles of Asheville, NC to ensure
that they had some familiarity with the area, with a sheet of color photographs representing three
stages of forest decline and explained that, without forest protection programs, high-elevation
spruce forests would all  decline to worst conditions (with severe tree mortality - an aerial survey
(Dull et al.  1988) had determined that in approximately one quarter of the area greater than 70%
of the standing trees were dead) and two potential forest protection programs. The first program
would protect the forests along road, and trail corridors spanning approximately 1/3 of the
ecosystem at risk.  This level of protection may be most appealing to recreational users.  The
second level of protection was for the entire ecosystem and may be most appealing to those who
value the continued existence of the entire ecosystem. Median household WTP was estimated to
be roughly $29 (in 2007  dollars) for the minimal program and $44 for the more extensive
program. Respondents were then asked to decompose their value for the extensive program into
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use, bequest, and existence values.  This resulted in values that represented components of 13%
use value, 30% bequest, 57% existence value (Table 4-7).
       Table 4-7.  Value Components for WTP for Extensive Protection Program for Southern
       Appalachian Spruce-Fir Forests.
Type of Value
Use
Bequest
Existence
Total
Proportion of WTP
0.13
0.30
0.57
1
Component Value
In $2007
$5.72
$13.20
$25.08
$44.00
       Another study by Jenkins, Sullivan, and Amacher (2002) estimated values for
recreational users of this resource. Households in the seven state Appalachian region (North
Carolina, South Carolina, West Virginia, Virginia, Tennessee, Kentucky, and Georgia) were
shown photographs taken in summer 1994 at permanent test plots. These photographs depicted
an unimpacted area (5% dead basal area), an area in beginning stage decline (30% dead basal
area) and severe decline (75% dead basal area) intended to represent a future scenario without a
forest protection program. The study estimated the mean WTP for forest protection programs at
$208 ($2007); multiplying across the population of the seven state area gives an aggregate
annual value of $3.4 billion for avoiding a significant decline in the health of high-elevation
spruce forests in the Southern Appalachian region (Table 4-8). This estimate is considerably
larger than the previous study.  The difference may be due to the closer proximity of the
respondents in the Jenkins et al.  study (the Kramer study radius of 500 miles includes
respondents as far away as Michigan, New York, Missouri, Louisiana and Florida).
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       Table 4-8. Summary of WTP Studies for Protection of Spruce Forests in the Southern
       Appalachians.
WTP in $ 2007
$29
$44
$208
$3.4b
Aggregation
Per household* for minimal program
Per household for extensive program
Per household for forest protection program
Jenkins estimate applied to the population
of the seven state** area included in the
study
Study
Kramer et al. (2003)

Jenkins (2002)

  * Households within a 500 mile radius of Asheville, NC
   **North Carolina, South Carolina, West Virginia, Virginia, Tennessee, Kentucky, and
      Georgia

       Forests in the northeastern United States also support and provide a wide variety of
valuable regulating services, including soil stabilization and erosion control, water regulation,
and climate regulation (Krieger, 2001). Forest vegetation plays an important role in maintaining
soils in order to reduce erosion, runoff, and sedimentation that can adversely impact surface
waters. In addition to protecting the quality of water in this way, forests also help store and help
regulate the quantities and temporal discharge patterns of water in watersheds. Forests also play
an important role in carbon sequestration at both regional and global scales. The total value of
these ecosystem services is very difficult to quantify.
       In some ecosystems where nitrogen is a growth limiting nutrient, there is the potential
that atmospheric deposition of nitrogen can increase biomass production of managed and
unmanaged systems with a consequent increase in carbon sequestration. The available evidence
(EPA,  2008) indicates N increases ecosystem carbon content in forested ecosystems, however
the magnitude of N stimulation of carbon sequestration is highly uncertain.  N effects on the
carbon budget of wetlands, grasslands and tundra is highly uncertain with respect to the direct
relationship between atmospheric nitrogen deposition and carbon sequestration. For all
ecosystem types there is the possibility of various unintended consequences, especially in non-
managed systems that occur as a result increased biomass production.   Consequently, data do
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not exist to adequately consider the potential changes of carbon sequestration associated with
atmospheric deposition of nitrogen.
4.4.4   Value Of Current Ecosystem Service Impairments Due To Terrestrial Acidification
        The REA Appendix 8 describes an analysis of ecosystem service impairments associated
with the impacts of terrestrial acidification on the forest product provisioning services from two
commercially important tree species on unmanaged forests - sugar maple and red spruce - that
are particularly sensitive to the effects of acidification.  Evidence of effects due to terrestrial
acidification is particularly  strong for these two species whose range includes the northeastern
U.S. where levels of nitrogen and sulfur deposition have historically been relatively high,
however more widespread impacts that include other tree species are also possible.   We
acknowledge that there may be some beneficial fertilization  effects of nitrogen deposition
however given the  complexity of the nitrogen cycle it is not  possible to quantify all those effects
here.  There is a detailed discussion of nitrogen fertilization effects in Chapter 3.
       In an exploratory study that is still under development we used data from the USFS
Forest Inventory and Analysis (FIA) database, to estimate an exposure-response relationship for
each species to measure the average negative effect of critical load exceedances (CLEs) of
nitrogen and sulfur deposition on annual tree growth. These estimated relationships  were then
applied  to sugar maple and  red spruce stocks in the Northeast and North central regions to
estimate the average percent increase in annual tree growth that would occur if all CLEs were
eliminated.  To estimate the aggregate-level forest market impacts of eliminating CLEs starting
in the year 2000, the tree-level growth adjustments were applied using the Forest and
Agricultural Sector Optimization Model  (FASOM), which is a dynamic optimization model of
the U.S. forest and agricultural sectors. The model results are reported as the present discounted
values of future welfare changes  in the forestry sector (in 5-year increments from 2000 to 2080)
due to increased tree growth. Summing over this 80-year period, the total present value of these
welfare  gains is $40.705 million  (in 2006 dollars, using a 4% discount rate). On an annualized
basis (at 4%), this is equivalent to $1.64 million per year. These estimates can  also be
interpreted as the current value of impairments to forest provisioning services provided by red
spruce and sugar maple due to acidification effects from nitrogen and sulfur. These results
should be considered very uncertain due to the pending revision of the exposure - response curve
and release of an updated version of the FASOM model. Referring back to the previous section it
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is apparent that cultural services supplied by these forests, including existence and recreational
use are a much larger category than the provisioning services estimated for these two species.

4.4.5   Aquatic Nutrient Enrichment
       Estuaries in the eastern United States are important for fish and shellfish production. The
estuaries are capable of supporting large stocks of resident commercial species, and they serve as
the breeding grounds and interim habitat for several migratory species (U.S. EPA, 2009). To
provide an indication of the magnitude of provisioning services associated with coastal fisheries,
from 2005 to 2007, the average value of total catch was $1.5 billion per year in 15 East Coast
states. It is not known, however, what percentage of this value is directly attributable to or
dependent upon the estuaries in these states. Based on commercial landings in Maryland and
Virginia, the values for three key species—blue crab, striped bass, and menhaden- totaled nearly
$69 million in 2007 in the Chesapeake Bay alone.
       Assessing how eutrophication in estuaries affects fishery resources requires bioeconomic
models (i.e., models that combine biological models offish population dynamics with economic
models describing fish harvesting and consumption decisions), but relatively few exist (Knowler,
2002). Kahn and Kemp (1985) estimated that a 50% decline in submerged aquatic vegetation
(SAV) from levels existing in the late  1970s (similar to current levels [Chesapeake Bay Program,
2008]) would decrease the net social benefits from striped bass by $16 million (in 2007 dollars).
In a separate analysis, Anderson (1989) modeled blue crab harvests under baseline conditions
and under conditions with "full restoration" of SAV. In equilibrium, the increase in annual
producer surplus and consumer surplus with full restoration of SAV was estimated to be $7.9
million (in 2007 dollars) or an 11% increase from current  service provision from blue crab alone.
Mistiaen et al. (2003) found that reductions in dissolved oxygen (DO) cause a statistically
significant reduction in commercial harvest and revenues crab harvests. For the Patuxent River
alone, a simulated reduction of DO from 5.6 to 4.0 mg/L was estimated to reduce crab harvests
by 49% and reduce total annual earnings in the fishery by  $275,000 (in 2007 dollars). While
these values do not quantify the increase in terms of atmospheric loadings alone, the estimated
20% loading to the Potomac River watershed (REA 5.2.4) from atmospheric deposition indicates
that the benefits apportioned to deposition are significant.
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       In addition, eutrophication in estuaries may also affect the demand for seafood. For
example, a well-publicized toxic pfiesteria bloom in the Maryland Eastern Shore in 1997 led to
an estimated $56 million (in 2007 dollars) in lost seafood sales for 360 seafood firms in
Maryland in the months following the outbreak (Lipton, 1999).  Surveys by Whitehead, Haab,
and Parsons (2003) and Parsons et al. (2006) indicated a reduction in consumer surplus due to
eutrophication-related fish kills ranging from $2 to $5 per seafood meal.8 As a result, they
estimated aggregate consumer surplus losses of $43 million to $84 million (in 2007 dollars) in
the month after a fish kill.
       As mentioned in the REA (5.2.1.3), estuaries in the eastern United States also provide an
important and substantial variety  of cultural ecosystem services, including water-based
recreational and aesthetic services. For example, FHWAR data  indicate that 4.8% of the
population in coastal states from North Carolina to Massachusetts participated in saltwater
fishing, with a total of 26 million saltwater fishing days in 2006 (U.S. DOT, 2007). Based on
estimates in Section 5.2.1.3 of the REA, total recreational consumer surplus value from these
saltwater fishing days was approximately $1.3 billion (in 2007 dollars). Recreational
participation estimates for several other coastal recreational activities are  also available for
1999-2000 from the NSRE. Almost 6 million individuals participated in motorboating in coastal
states from North Carolina to Massachusetts.  Again, based on analysis in  the REA, the aggregate
value of these coastal motorboating outings was $2 billion per year.  Almost 7 million people
participated in birdwatching, for a total of almost  175  million days per year, and more than 3
million participated in visits to nonbeach coastal waterside areas, for a total of more than 35
million days per year (Table 4-9).
       Estuaries and marshes have the potential to support a wide range of regulating services,
including climate, biological, and water regulation; pollution detoxification; erosion prevention;
and protection against natural hazards (MEA, 2005c). The relative lack of empirical models and
valuation studies imposes obstacles to the estimation of ecosystem services affected by nitrogen
deposition.  While atmospheric deposition contributes to eutrophication there is uncertainty in
separating the effects of atmospheric nitrogen from nitrogen reaching the estuaries from  many
other sources.
 Surprisingly, these estimates were not sensitive to whether the fish kill was described as major or minor or to the
different types of information included in the survey.

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Table 4-9.  Summary of Values for Current Levels of Services and Changes in
Service Levels in $2007.
Ecosystem Service
Total Catch -
Commercial Fishing

Change in Ecosystem
Service
50% decline in SAV
Full restoration SAV
0.4% mg/L decrease DO
HAB

Ecosystem Service
Saltwater fishing
Motorboating
Bird watching
Non-beach coastal visits
Area or Population
Affected
14 east coast states
MD/VA

Chesapeake Bay
Chesapeake Bay
Patuxent River
1 997 MD eastern
shore


4.6% pop. MA-NC
6 million
7 million
3 million
Value
($2007)
$1.5b/yr
$69 m/yr
Value of
Change
($2007)
1 $16 m/yr
t $ 8 m/yr
| $275 th/yr
| $56 m
I $43-84 m
Participation
Days
26m days

175 mdays
35 mdays
Species

blue crab, striped bass,
menhaden

striped bass
blue crab
^ 49% blue crab harvest
loss to seafood industry
sustained loss over months
Value ($2007)
$1.3b/yr
2b/yr


Note: Down arrows indicate a decrease in value of the magnitude specified: up arrows
indicate an increase in value of the magnitude specified.
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4.4.6 Value of Current Ecosystem Service Impairments Due to Aquatic Nutrient Enrichment
       The aquatic nutrient enrichment case study relied on the NOAA Eutrophication Index as
the indicator, which includes dissolved oxygen, HABs, loss of SAV and loss of water clarity.
There are methods available to link some of the components to ecosystem services, most notably
loss of SAV and reductions in DO. The REA analysis estimates the change in several ecosystem
services including recreational fishing, boating, beach use, aesthetic services and nonuse
services. The REA focuses on two major East Coast estuaries - the Chesapeake Bay and the
Neuse River. Both estuaries receive between 20%-30% percent of their annual nitrogen loadings
through atmospheric deposition and both are showing symptoms of eutrophication. The analysis
uses and adapts results from several existing studies to approximate effects on several ecosystem
services, including commercial fishing, recreation, aesthetic enjoyment, and nonuse values. For
example, it is estimated that atmospheric nitrogen decreases the annual benefits of recreational
fishing, boating, and beach use in the Chesapeake Bay by $43-$217 million, $3-8 million, and
$124 million respectively, and reduces annual aesthetic benefits to nearshore residents by $39-
102 million (Table 4-10). In the Neuse River, the value of annual commercial crab fishing
services would be between $0.1-1 million higher without the contribution of atmospheric
nitrogen, and recreation fishing services in the larger Albermarle Pamlico Sound estuary system
(which includes the Neuse) would be $1-8 million greater per year.
       Table 4-10.  Summary of Annual Damages to Services due to Atmospheric Loading.
Ecosystem Service
Recreational Saltwater Fishing

Beach Use
Boating

Commercial Crab Fishing
Annual Value ($2007)
$43-217 b
$1-8 m
$39-102 m
$3-8 m

$0.1-1 m
Waterbody Affected
Chesapeake Bay
Albemarle Pamlico Sound
Chesapeake Bay
Chesapeake Bay

Neuse River
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4.4.7  Terrestrial Nutrient Enrichment
       For the purposes of the following section nutrient enrichment refers only to that due to
NOx deposition. Additionally these sections focus on the detrimental effects of that deposition.
Staff acknowledges that NOX deposition in managed terrestrial ecosystems has a beneficial
effect, specifically increased growth (a fertilization effect).  These effects are discussed in detail
in Chapter 3 of this document.
       The ecosystem service impacts of terrestrial nutrient enrichment in unmanaged ecosystems
include primarily cultural and regulating services. In CSS areas, concerns focus on a decline in
CSS and an increase in nonnative grasses and other species, impacts on the viability of threatened
and endangered species associated with CSS, and an increase in fire frequency. Changes in MCF
include changes in habitat suitability and increased tree mortality, increased fire intensity, and a
change in the forest's nutrient cycling that may affect surface water quality through nitrate
leaching  (EPA, 2008).
       The terrestrial nutrient enrichment case study relies  on benchmark deposition levels for
various species and ecosystems as indicators of ecosystem response.  While it would be expected
that deposition above those levels would have deleterious effects on the provision  of ecosystem
services in those areas, at this time it is possible only to describe the magnitude of the some of
the services currently being provided.  Methods are not yet  available to allow estimation of
changes in services due to nitrogen deposition.  For the  purposes of the following sections
nutrient enrichment refers only to that  due to NOX deposition. Additionally these sections focus
on the detrimental  effects of that deposition.  Staff acknowledges that a certain amount of NOX
deposition in managed terrestrial ecosystems has  a beneficial  effect, specifically increased
growth (a fertilization effect). However no attempt has been made to quantify those beneficial
effects since this document and preceding analyses are focused on unmanaged sensitive
ecosystems.
       The value that California residents and the U.S. population as a whole place on CSS and
MCF habitats is reflected in the various federal, state, and local government measures that have
been put in place to protect these habitats. Threatened and endangered species are  protected by
the Endangered Species Act. The State of California passed the Natural Communities
Conservation Planning Program (NCCP) in 1991, and CSS  was the first habitat identified for
protection under the program (see www.dfg.ca.gov/habcon/nccp). It is estimated that only 10 -
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15% of the original extent of CSS habitat remains (NPS.gov/cabr/naturescience/coastal-sage-
scrub-and-southern-chaparrel-communities.htm). Private organizations such as The Nature
Conservancy, the Audubon Society, and local land trusts also protect and restore CSS and MCF
habitat.
       CSS and MCF are found in numerous recreation areas in California. Three national parks
and monuments in California contain CSS, including Cabrillo National Monument, Channel
Islands National Park, and Santa Monica National Recreation Area. All three parks showcase
CSS habitat with educational programs and information provided to visitors, guided hikes, and
research projects focused on understanding and preserving CSS. Over a million visitors traveled
through these three parks in 2008. MCF is highlighted in Sequoia and Kings Canyon National
Park, Yosemite National Park, and Lassen Volcanic National Park, where more than 5 million
people visited in 2008.
       The 2006 FHWAR for California (DOT, 2007) reports on the number of individuals
involved in fishing, hunting, and wildlife viewing in California. Millions of people are involved
in just these three activities each year. The  quality of these trips depends in part on the health of
the ecosystems  and their ability to support the diversity of plants and animals found in important
habitats found in CSS or MCF ecosystems  and the parks associated with those ecosystems.
Based on analyses in Section 5.3.1.3 of the REA (U.S.EPA, 2009), average values of the total
benefits in 2006 from fishing, hunting, and wildlife viewing away from home in California were
approximately $947 million, $169 million,  and $3.59 billion, respectively (Table 4-11). In
addition, data from California State Parks (2003) indicate that in 2002, 68.7% of adult residents
participated in trail hiking for an average of 24.1 days per year. The analyses in the REA
(U.S.EPA, 2009) indicate that the aggregate annual benefit for California residents from trail
hiking  in 2007 was $11.59 billion. It is not currently possible to quantify the loss in value of
services due to  nitrogen deposition as those losses are already reflected in the estimates of the
contemporaneous total value of these recreational activities.  Restoration of services through
decreases in nitrogen deposition would likely increase the total value of recreational services.
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       Table 4-11. Summary of Current Levels of Ecosystem Services.
Activity


Trail Hiking

Fishing
Wildlife
Viewing
Hunting
Participation


68. 7% of C A
population
1.7m
6.2m

0.28m
#of
days/yr

453m

19m
45m

3.3m
Average
WTP

$25.59

$48.86
$79.81

$50.10
Annual Aggregate
Value ($2007 in
millions)
115,900

947
3,600

169
       Sources: 2006 FHWAR for California (DOI, 2007), California State Parks (2003),
       Kaval and Loomis (2003)
       CSS and MCF are home to a number of important and rare species and habitat types. CSS
displays richness in biodiversity with more than 550 herbaceous annual and perennial species. Of
these herbs, nearly half are endangered, sensitive, or of special status (Burger et al, 2003).
Additionally, avian, arthropod, herpetofauna, and mammalian species live in CSS habitat or use
the habitat for breeding or foraging. Communities of CSS are home to three important federally
endangered species: the Quino checker spot butterfly, the kangaroo rat and the California
gnatcatcher. MCF is home to one federally endangered species (mountain yellow-legged frog)
and a number of state-level sensitive species. The Audubon Society lists 28 important bird areas
in CSS habitat and at least 5 in MCF in California (http://ca.audubon.org/iba/index.shtml).9
       The terrestrial enrichment case study in Section 5.3.1.3 of the REA and  Section 3.3.5 of
the ISA identified fire regulation as a service that could be affected by nutrient enrichment of the
CSS and MCF ecosystems by encouraging growth of more flammable grasses, increasing fuel
loads, and altering the fire cycle. Over the 5-year period from 2004 to 2008, Southern California
experienced,  on average,  over 4,000 fires per year burning, on average, over 400,000 acres per
year (National Association of State Foresters [NASF], 2009). It is not possible at this time to
quantify the contribution of nitrogen deposition, among many other factors, to increased fire risk.
       The CSS and MCF were selected as case  studies for terrestrial enrichment because of the
potential that these areas  could be adversely affected by excessive N deposition. To date, the
' Important Bird Areas are sites that provide essential habitat for one or more species of bird.
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detailed studies needed to identify the magnitude of the adverse impacts due to N deposition
have not been completed. Based on available data, this report provides a qualitative discussion of
the services offered by CSS and MCF and a sense of the scale of benefits associated with these
services. California is famous for its recreational opportunities and beautiful landscapes. CSS
and MCF are an integral part of the California landscape, and together the ranges of these
habitats include the densely populated and valuable coastline and the mountain areas. Through
recreation and scenic value, these habitats affect the lives of millions of California residents and
tourists. Numerous threatened and endangered species at both the state and federal levels reside
in CSS and MCF. Both habitats may play an important role in wildfire frequency and intensity,
an extremely important problem for California. The potentially high value of the ecosystem
services provided by CSS and MCF justify careful attention to the long-term viability of these
habitats.
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National Oceanic and Atmospheric Administration (NOAA). "Recreational Fisheries."
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New Hampshire Department of Environmental Services, Prepared by Robert H. Estabrook.
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       FINAL, September, 2004.
New York State, Department of Environmental Conservation Division of Water, Bureau of
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Nilsson, J. and Grennfelt, P. (ed.):1988, Critical loads of acidity for sulfur and nitrogen, Nordic
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Olson, Mancur.  The Logic of Collective Action : Public Goods and the Theory of Groups
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Parsons, G., A.O. Morgan, J.C. Whitehead, and T.C. Haab. 2006. "The Welfare Effects of
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5      CO-PROTECTION FOR OTHER EFFECTS POTENTIALLY
       AFFORDED BY AN AQUATIC ACIDIFICATION STANDARD

       This chapter focuses on the co-protection that a standard focused on aquatic acidification
might afford for other deposition related ecological effects, including terrestrial acidification,
terrestrial nutrient enrichment, and estuarine eutrophication.
5.1    POTENTIAL CO-PROTECTION FOR TERRESTRIAL ACIDIFICATION
       To understand the level of co-protection an aquatic acidification standard for oxides of
nitrogen and sulfur might afford for terrestrial ecosystems, an analysis was conducted to compare
the critical acid loads for aquatic and terrestrial components of watersheds in the eastern United
States. Aquatic critical acid loads are an integrated function of the chemistry of runoff from
stream and lake waters, and the biogeochemical processes that occur within the aquatic and
terrestrial components of the entire watershed.  Terrestrial critical acid loads, however, are
largely determined by the conditions and processes that occur in the root zone of the soil profile
of the terrestrial systems  of a watershed. Therefore, it is possible to have different critical acid
load values for aquatic and terrestrial ecosystems within the same watershed.
       For the comparative analysis of aquatic and terrestrial critical  acid loads, aquatic critical
acid loads were selected based on an acid neutralizing capacity (ANC) of 50ueq/L, and
weretaken directly from the Risk and Exposure Assessment (REA). The 50ueq/L ANC value
was one of three example values modeled in the REA for aquatic acidification.  The terrestrial
critical acid loads in this  comparative analysis were selected to protect for either a terrestrial base
cation to aluminum molar ratio (Bc:Al) of 1.2 or 10.0.  The Bc:Al ratio of 10.0 would be
relatively more protective, as it provides greater protection against the impacts of acidification on
cation availability and aluminum toxicity in the soil solution. The terrestrial critical loads were
calculated using the Simple Mass Balance (8MB) method outlined in the REA and input values
averaged across the area  of each watershed.
       Aquatic and terrestrial critical acid loads were compared in 16 watersheds from each of
the two aquatic acidification case study areas, the Adirondacks and the Shenandoah, identified in
the REA.  For each case study area, four watersheds were randomly selected from each of the
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four aquatic acidification sensitivity classes reported in the REA.  Those four sensitivity classes
are "highly sensitive", "moderately sensitive", "low sensitivity", and "not sensitive". In order
for a watershed to be classified as one of these four classes, it had to contain at least one lake or
stream with that sensitivity class designation. The Adirondacks case study area contained
watersheds representing all four sensitivity classes, and the 16 watersheds that were selected for
the analysis contained a total of 29 lakes. However, in the Shenandoah case study area, there
were a limited number of watersheds in the "low" and "not sensitive" classes. Therefore, only
one of the 16 randomly selected watersheds contained a "low" and a "not sensitive" stream. In
total, there were 20 streams located in the 16 Shenandoah watersheds selected for the
comparative analysis. In each of the 32 watersheds (16 Adirondacks plus 16 Shenandoah), the
terrestrial critical acid loads were calculated as a single value for the entire watershed. These
terrestrial critical acid loads were then compared to the aquatic critical acid loads for the lakes
and streams within each watershed to determine whether the aquatic or terrestrial critical acid
load provided greater protection against acidifying nitrogen and sulfur deposition.  Appendix A
provides a full description of the methods and results of this comparative analysis.
       Results of the comparison between the example aquatic critical acid load (ANC = 50
ueq/L) and the terrestrial critical acid loads (Bc:Al 1.2 and 10.0) for the 32 watersheds are
presented in Tables 5.1 and 5.2. In the 16 Adirondack watersheds, 13 of the 29 lakes had aquatic
critical acid loads that were lower (more protective) than the terrestrial critical acid loads when a
Be: Al ratio of 10.0 was used.  Based on terrestrial critical acid loads determined with a Be: Al
ratio of 1.2, 21 of the 29 lakes in the Adirondacks had aquatic critical acid loads lower than the
terrestrial critical acid loads.  More importantly, for the terrestrial critical acid loads determined
with a Be: Al ratio of 10.0,  13 of the 16 lakes in the Adirondacks classified as "highly" and
"moderately" sensitive to acidification had aquatic critical acid loads lower than the terrestrial
critical acid loads, and all 16  lakes in these two sensitivity classes had critical acid loads lower
than the terrestrial loads determined with a Bc:Al of 1.2  The watersheds within the Shenandoah
region showed similar results (Tables 5.1 and 5.2).
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Table 5-1. Results of the comparison of lake and stream aquatic critical loads (ANC of 50
jieq/L) to terrestrial critical loads.
(Bc:Al molar ratios of 10.0 in soil solution) calculated for the full watershed in each of the 16 watersheds in the
Adirondacks and Shenandoah case study areas.  The tabular results show the number of times the aquatic
acidification critical load would provide more protection than the terrestrial acidification critical load.
Case Study Area
Adirondacks
Shenandoh
Watershed Sensitivity to Aquatic Acidification
Highly Sensitive
7 of 7
14 of 14
Moderately Sensitive
6 of 9
5 of 5
Low Sensitivity
Oof 7
Oof 1
Not Sensitive
Oof 6
Oof 1
Table 5-2. Results of the comparison of lake and stream aquatic critical loads (ANC of 50
ueq/L) to terrestrial critical loads.
(Bc:Al molar ratios of 1.2 in soil solution) calculated for the full watershed in each of the 16 watersheds in the
Adirondacks and Shenandoah Case Study Areas. The tabular results show the number of times the aquatic
acidification critical load would provide more protection than the terrestrial acidification critical load.
Case Study Area
Adirondacks
Shenandoh
Watershed Sensitivity to Aquatic Acidification
Highly Sensitive
7 of 7
14 of 14
Moderately Sensitive
9 of 9
5 of 5
Low Sensitivity
5 of 7
Oof 1
Not Sensitive
Oof 6
Oof 1
       In summary, terrestrial and aquatic critical acid loads were compared for watersheds in
the Adirondack and Shenandoah case study areas. Results indicated that, in general, the aquatic
critical acid loads were lower and therefore offered greater protection to the watershed than did
the terrestrial critical acid loads. In situations where the terrestrial critical acid loads were lower
(i.e., more protective) than the aquatic critical acid loads, the lakes or streams in the watershed
were often rated as having "low sensitivity" or "not sensitive" to acidifying nitrogen and sulfur
deposition.  Conversely, when the waterbodies were more sensitive to deposition ("highly
sensitive" or "moderately sensitive"), the aquatic critical acid loads generally provided a greater
level of protection against acidifying nitrogen and sulfur deposition in the watershed.  It is
uncertain whether these results would be consistent for the rest of the country.
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5.2    POTENTIAL CO-PROTECTION FOR TERRESTRIAL NUTRIENT
ENRICHMENT
       Whereas critical loads have been modeled for aquatic acidification protection, for
terrestrial nutrient enrichment protection, there are only empirical benchmarks developed from
field studies of specific ecosystems or species. To understand the level of co-protection an
aquatic acidification standard for oxides of nitrogen and sulfur might afford for nutrient over-
enrichment effects to terrestrial ecosystems, a critical load developed for a lake in an ecoregion
can be compared to nitrogen deposition benchmarks found in the scientific literature. Figure 5.1
below summarizes the terrestrial nutrient enrichment effects summarized in the REA.
       For each depositional load that is considered for aquatic acidification, it can be compared
against the chart in Figure 5.1 to understand the level of protection offered in individual parts of
the country where these studies were conducted.  For example, if a maximum nitrogen
depositional load was selected for California of 80 meq/(m2-yr) or ~11 kg N/(ha-yr),  this could
be compared directly to the benchmarks in Figure 5.1 that describe California ecosystems
(meq/(m2-yr) divided by 7.14 equals kg/(ha-yr).  Comparing this  maximum  nitrogen  deposition
number to the benchmarks in Figure 5.1, shows that the depositional load would provide some
protection against leaching in the San Bernardino Mountains (11-40 kg/(ha-yr)), but would have
to be lower to protect California coastal sage scrub in Southern California (3.1-3.3 kg/(ha-yr)),
and lichens in mixed conifer forests (3.1-5.3 kg/(ha-yr)).
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              Rocky Mountain alpine lakes: shift in diatom community dominance (Baron, 2006)


               •  Southern California: N growth requirement threshold  (Wood et al., 2006)
               •  San Bernardino Mountains and Sierra Nevada Mountains: acidophytic lichen
                  decline in MCF (Fenn et al., 2008)

              •  Eastern Rocky Mountain Slope: low carbon:nitrogen; low lignin:nitrogen (Baron et
                 al., 2000)
              •  Eastern Rocky Mountain Slope: increased foliar nitrogen; increased mineralization
                 (Baron et al., 2000)

                 •   San Bernardino Mountains and Sierra Nevada Mountains: shift from acidophytic
                    to neutral or nitrogen-tolerant lichen in MCF (Fenn et al., 2008)
                 •   Minnesota grasslands: decreased plant species (Clark and Tilman, 2008)

                   •  Northeast U.S.: NO3 leaching (Aber et al., 2003)
                        Bay Area, CA: Increased cover of nonnative grasses; decreased native
                        grasses (Weiss, 1999)

                        San Bernardino Mountains and Sierra Nevada Mountains: loss of acidophytic
                        lichen in MCF (Fenn et al., 2008)
                        Southern California: shift in mycorrhizal species in CSS (Egerton-Warburton
                        and Allen, 2000)
                        Southern California: shift from native species to invasive grasses in CSS (Allen,
                        2008)*
                        •   San Bernardino Mountains: high dissolved organic nitrogen (Meixnerand
                            Fenn, 2004)
                        •   San Bernardino Mountains: nitrogen saturation (Meixner and Fenn, 2004)
                             Increased nitrogen in lichen (Fenn et al., 2007)
                                        MCF: NO3 leaching (Fenn et al., 2008)
                                        MCF: 25% decrease in fine-root biomass (Fenn et al., 2008)

                                         Southern California: NO3  leaching (Fenn et al., 2003)
                                         Southern California: high foliar nitrogen (Bytnerowicz and
                                         Fenn, 1996)
                                         Los Angeles Basin, California: High NO emissions
                                         (Bytnerowicz and Fenn, 1996)
                                                        •  Fraser Experimental Forest, CO:
                                                          increased foliar nitrogen; increased
                                                          mineralization (Rueth et al., 2003)**
                 10
15       20       25        30       35
Nitrogen Deposition,  kg/ha/yr
40
45
" Personal communication, 2008, Also referenced in Bobbink et. al. 2010, Ecological Applications,20(1): 30-69 and USDS FS, 2010,
http:/Avww.nrs.fs.f^.u^l^n_alrj«ater^ctean_Miater/critlcal_lDadsflocal-rasoijtcas/
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5.3 POTENTIAL CO-PROTECTION FOR AQUATIC NUTRIENT ENRICHMENT
       The RE A found that deposition of reactive nitrogen contributed to eutrophication of
estuaries; however, it was also noted that atmospheric deposition of nitrogen is only part of the
total nitrogen load to the estuaries. This makes it difficult to understand the co-protection that an
aquatic acidification standard for oxides of nitrogen and sulfur might afford.  One way to
approach this issue is to assume any reduction in atmospheric nitrogen load to an enriched
estuary will be a benefit. A second approach would be to see how far the reduction in nitrogen
likely to result from an aquatic acidification standard would go towards meeting the total
nitrogen load reduction goals set by the estuary programs.
       As described in the REA, the Chesapeake Bay is one national estuary that is suffering
from eutrophication.  In issuing his Executive Order on the Chesapeake Bay (EO 13508),
President Obama recognized that the Bay watershed is one of our nation's greatest treasures and
must be protected and restored. To that end, EPA is proposing a nitrogen total maximum daily
load (TMDL) for the Chesapeake Bay. The TMDL will contain a specific air allocation for
nitrogen deposition. The allocations that were provided to the states included assumptions that
air deposition levels of nitrogen would be reduced to 14.9 million pounds per year  to the tidal
waters and to 323 million pounds to the watershed by the year 2020.  According to the
Chesapeake Bay Program Office, the tidal waters have a surface area of 4,479 square miles and
the watershed is 64,216 square miles.  This means that in 2020, the TMDL currently calls for
nitrogen deposition levels to the combined bay and watershed to be reduced to 337.9 million
pounds/68,695 square miles/yr, which is equivalent to 8.6 kg/ha/yr or 61 meq/m2/yr. As in
Section 5.2, if we use as an example a maximum nitrogen depositional load of 80 meq/m2/yr or
~11 kg N/ha/yr, this number would not meet the Chesapeake Bay TMDL as currently
envisioned.
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5.4    REFERENCES

Aber JD; Goodale CL; Ollinger SV; Smith ML; Magill AH; Martin ME; Hallett RA; Stoddard
       JL. (2003). Is nitrogen deposition altering the nitrogen status of northeastern forests?
       Bioscience, 53, 375-389.
Allen EB; Rao LE; Steers RJ. (2008). Impacts of atmospheric nitrogen deposition on vegetation
       and soils at Joshua Tree National Park. In: Webb RH; Fenstermaker LF; Heaton JS;
       Hughson DL; McDonald EV; Miller DM (Eds.), The Mojave Desert: Ecosystem
       processes and sustainability. Reno, NV: University of Nevada Press.
Baron, J.S., H.M. Rueth, A.M. Wolfe, K.R. Nydick, E.J. Allstott, J.T. Minear, and B.
       Moraska.(2000). Ecosystem responses to nitrogen deposition in the Colorado Front
       Range. Ecosystems 3:352-368.
Baron JS. (2006). Hindcasting nitrogen deposition to determine ecological critical load. Ecol
       Appl,  16, 433-439.
Bytnerowicz,  A., and M.E. Fenn. (1996). Nitrogen deposition in California forests:  a review
       Environmental Pollution 92(2): 127-146.
Clark CM; Tilman D. (2008). Loss of plant species after chronic low-level nitrogen deposition to
       prairie grasslands. Nature, 451, 712-715.
Egerton-Warburton, L.M., and E.B. Allen. (2000). Shifts in arbuscular mycorrhizal communities
       along  an anthropogenic nitrogen deposition gradient. Ecological Applications 70(2):484-
       496.
Executive Order 13508. 2009. Chesapeake Bay Protection and Restoration. Federal Register Vol
       74 No 93: 23099-23104.
Fenn, M.E., J.W. Baron, E.B. Allen, H.M. Rueth, K.R. Nydick, L. Geiser, W.D. Bowen, J.O.
       Sickman, T. Meixner, D.W. Johnson, and P. Neitlich. (2003). Ecological effects of
       nitrogen deposition in the western United States. Bioscience 53(4):404-420.
Fenn, M.E., L. Geiser, R. Bachman, T.J. Blubaugh, and A. Bytnerowicz. (2007). Atmospheric
       deposition inputs and effects on lichen chemistry and indicator species in the Columbia
       River  Gorge, USA. Environmental Pollution 146:77-91.
Fenn, M.E., S. Jovan, F. Yuan, L. Geiser,  T. Meixner, and B.S. Gimeno. (2008). Empirical and
       simulated critical loads for nitrogen deposition in California mixed conifer forests.
       Environmental Pollution 755(3):492-511.
Meixner, T., and M. Fenn. (2004). Biogeochemical budgets in a Mediterranean catchment with
       high rates of atmospheric N deposition -  Importance of scale and temporal asynchrony.
       Biogeochemistry 70:3 31 -3 5 6.
Rueth,  H.M.,  J.S. Baron, and E.J.  Allstott. 2003.  Responses of Engelmann spruce forests to
       nitrogen fertilization in the Colorado Rocky Mountains. Ecological Applications 13:664-
       673.
U.S. EPA (Environmental Protection Agency). 2009. Risk and Exposure Assessment for Review
       of the  Secondary National Ambient Air Quality Standards for Oxides of Nitrogen and
       Sulfur. U.S. Environmental Protection Agency, Office of Research and Development,
       National Center for Environmental Assessment, Research Triangle Park, NC. September.
Weiss, S.B. 1999. Cars, cows, and checkerspot butterflies: Nitrogen deposition and management
       of nutrient-poor grasslands for a threatened species. Conservation Biology 13:1476-
       1486.
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Wood YA; Meixner T; Shouse, PJ; Allen, EB. (2006). Altered ecohydrologic response drives
       native Shrub loss under conditions of elevated nitrogen deposition. J Environ Qual, 35,
       76-92.
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6       ADDRESSING THE ADEQUACY OF THE CURRENT STANDARDS

       Based on the information in Chapters 3 and 4, we conclude that there is support in the
available effects-based evidence for consideration of secondary standards for oxides of nitrogen
and sulfur that are protective against adverse ecological effects associated with deposition of
oxides of nitrogen and sulfur to sensitive ecosystems.  Having reached this general conclusion,
we then to the extent possible evaluate the adequacy of the current standards for oxides of
nitrogen and sulfur by considering to what degree risks to sensitive ecosystems would be
expected to  occur in areas that meet the current standards. Staff conclusions regarding the
adequacy of the current standards are based on the available ecological effects, exposure and
risk-based evidence.  In evaluating the strength of this information, staff have taken into account
the uncertainties and limitations in the scientific evidence.  This chapter addresses key policy
relevant questions that inform our determination regarding the adequacy of the structure and
levels of the current secondary standards. The chapter begins with a discussion of the structure
of the current standards, followed by a presentation of information on recent air quality relative
to the existing standards, recent deposition levels of oxides of nitrogen and sulfur, evaluation of
recent deposition levels relative to levels where adverse ecological effects have been observed,
and a set of conclusions regarding the adequacy of the current structure and levels of the
standards.   Acidification occurs over extended periods and the ability of both terrestrial and
aquatic systems to recover is dependent upon not only the decrease in acidic deposition, but the
ability of these ecosystems to generate cations needed for nutrients and base cation supply.  As a
result, given the same decrease in deposition, ecosystems with high levels of base cation
replacement will recover faster than those with low levels.

6.1    APPROPRIATENESS OF THE CURRENT STANDARD
       The  current secondary oxides of nitrogen and sulfur standards are intended to protect
against adverse effects to public welfare. For oxides  of nitrogen, the current secondary standard
was set identical to the primary standard1, e.g. an annual standard set for NC>2 to protect against
adverse effects on vegetation from direct exposure to ambient oxides of nitrogen.  For oxides of
1 The current primary NO2 standard has recently been changed to the 3 year average of the 98thpercentile of the
annual distribution of the 1 hour daily maximum of the concentration of NO2. The current secondary standard
remains as it was set in 1971.
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sulfur, the current secondary standard is a 3-hour standard intended to provide protection for
plants from the direct foliar damage associated with atmospheric concentrations of SC>2. It is
appropriate in this review to consider whether the current standards are adequate to protect
against the direct effects on vegetation resulting from ambient NC>2 and 862 which were the
basis for the current secondary standards. The ISA concluded that there was sufficient evidence
to infer a causal relationship between exposure to 862, NO, NC>2 and PAN and injury to
vegetation. Additional research on acute foliar injury has been limited and there is no evidence
to suggest foliar injury below the levels of the current secondary standards for oxides of nitrogen
and sulfur. There is sufficient evidence to suggest that the levels of the current standards are
likely adequate to protect against direct phytotoxic effects.
       The ISA however, has established that the major effects of concern for this review of the
oxides of nitrogen and  sulfur standards are associated with deposition of N and S caused by
atmospheric concentrations of oxides of nitrogen and sulfur (see Chapter 3).  As discussed in the
following sections, the current standards are not directed toward depositional effects, and none of
the elements of the current NAAQS - indicator, form, averaging time, and level - are suited for
addressing the effects of N and S deposition. Thus,  by using atmospheric NC>2 and SC>2
concentrations as indicators, the current  standards address only a fraction of total atmospheric
oxides of nitrogen and  sulfur, and do not take into account the effects from deposition of total
atmospheric oxides of nitrogen and sulfur. By addressing short-term concentrations, the current
862  standard, while protective against direct foliar effects from gaseous oxides of sulfur, does
not take into account the findings of effects in the ISA, which notes the relationship between
annual deposition of S  and acidification effects which are likely to be more severe and
widespread than phytotoxic effects under current ambient conditions, and include effects from
long term deposition as well as short term. Acidification is a process which occurs over time, as
the ability of an aquatic system to counteract acidic  inputs is reduced as natural buffers are used
more rapidly than they can be replaced through geologic weathering. The relevant period of
exposure for ecosystems is therefore not the exposures captured in the short averaging time of
the current 862 standard.
       The levels of the current standards also are not well suited to dealing with deposition-
based effects of oxides of nitrogen and sulfur.  Current standards are specified as allowable
single atmospheric concentration levels for NC>2 or 862. This type of structure does  not take into

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account variability in the atmospheric and ecological factors that may alter the effects of oxides
of nitrogen and sulfur on public welfare.  Consistent with section 108 of the CAA, the ISA
includes in the air quality criteria consideration of how these variable factors impact the effects
of ambient oxides of nitrogen and sulfur on public welfare. See CAA section 108 (a)(2)(A)
requiring air quality criteria to include information on "those variable factors (including
atmospheric conditions) which of themselves or in combination with other factors may alter the
effects on public health or welfare of such air pollutant." Secondary standards are intended to
address a wide variety of effects occurring in different types of environments and ecosystems.
Ecosystems are not uniformly distributed either spatially or temporally in their sensitivity to air
pollution.  Therefore, failure to account for the major determinants of variability, including
geological and soil characteristics related to the sensitivity to acidification as well as atmospheric
and landscape characteristics that govern rates of deposition, may lead to standards that do not
provide requisite levels of protection across ecosystems. Finally, given the mismatch of all of
the other elements of the current secondary NAAQS with deposition-based effects, the  form of
those standards will also be mismatched.
       Because all areas of the U.S. are in attainment with the current NC>2 and SC>2 standards,  it
is possible to evaluate current conditions, and evaluate the impact on public welfare from the
current effects on ecosystems from oxides of nitrogen and sulfur deposition in areas that attain
the current standards that use NC>2 and SC>2 as indicators. In addition, this chapter qualitatively
addresses  the adequacy of the structures  of the existing  standards relative to ecologically relevant
standards  for oxides of nitrogen and sulfur, and sets up arguments for developing an ecologically
relevant structure for the standards as described in chapter 7.

6.2     STRUCTURES OF THE CURRENT OXIDES OF NITROGEN AND SULFUR
       SECONDARY STANDARDS AND RELEVANT ECOLOGICAL INDICATORS
       OF PUBLIC WELFARE EFFECTS
       The current secondary standard for NOX, set in 1971, using NC>2 as the atmospheric
indicator,  is 0.053 parts per million (ppm) (100 micrograms per cubic meter of air [|ig/m3]),
annual arithmetic average, calculated as the arithmetic mean of the 1-hour NC>2 concentrations.
This standard was selected to provide protection to the public welfare against acute injury to
vegetation from direct exposure and resulting phytoxicity.  During the last review of the oxides
of nitrogen standards,  impacts associated with chronic acidification and eutrophication  from
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oxides of nitrogen deposition were acknowledged, but the relationships between atmospheric
concentrations of oxides of nitrogen and levels of acidification and eutrophication and associated
welfare impacts were determined to be too uncertain to be useful as a basis for setting a national
secondary standard (U.S. EPA ,1995).
       The current secondary standard for oxides of sulfur, set in 1971, uses 862 as the
atmospheric indicator, is a 3-hour average of 0.5 ppm, not to be exceeded more than once per
year.  This standard was selected to provide protection to the public welfare against acute injury
to vegetation.  In the last review of the oxides of sulfur secondary standard, impacts associated
with chronic acidification were acknowledged, but the relationships between atmospheric
concentrations of oxides of sulfur and levels of acidification, along with the complex interactions
between oxides of nitrogen and sulfur in acidification processes, were cited as critical
uncertainties which made the setting of secondary NAAQS  to protect against acidification
inappropriate at that time (U.S. EPA, 1982).
       In the previous separate reviews of the oxides of nitrogen and sulfur secondary standards,
EPA acknowledged in each review the additional impacts of oxides of nitrogen and sulfur on
public welfare through the longer term impact of the pollutants once deposited to ecosystems.
However, the previous reviews cited numerous uncertainties as the basis for not directly
addressing those impacts in the setting of secondary standards.  In addition, these previous
reviews did not consider the  common pathways of impact for both nitrogen and sulfur acting on
the same ecosystem endpoints.
       Three issues arise that call into question the ecological relevance of the current structure
of the secondary standards for oxides of nitrogen and sulfur. One issue is the exposure period
that is relevant for ecosystem impacts. The majority of deposition related impacts are associated
with depositional loads that occur over periods of months to years. This differs significantly
from exposures associated with hourly concentrations of NC>2 and SC>2 as measured by the
current standards. Even though the NC>2 standard uses an annual average of NC>2, it is focused on
the annual average of 1-hour NC>2 concentrations, rather than on a cumulative metric or an
averaging metric based on daily or monthly averages.  A  second issue is the choice of
atmospheric indicators. NC>2  and  862 are used as the component of oxides of nitrogen and sulfur
that are measured, but they do not provide a complete link to the direct effects on ecosystems
from deposition of oxides of nitrogen and sulfur as they do  not capture all relevant chemical

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species of oxidized nitrogen and oxidized sulfur that contribute to deposition. The ISA provides
evidence that deposition related effects are linked with total nitrogen and total sulfur deposition,
and thus all forms of oxidized nitrogen and oxidized sulfur that are deposited will contribute to
effects on ecosystems.  This suggests that more comprehensive atmospheric indicators should be
considered in designing ecologically relevant standards.  Further discussions of the need for
more ecologically relevant atmospheric indicators as well as the relative contributions to
deposition from various species of oxides of nitrogen and sulfur can be in found in Chapter 2.
The third issue is that the current standards reflect separate assessments of the two individual
pollutants, NC>2 and SC>2, rather than assessing the joint impacts of deposition to ecosystems,
recognizing the role that each pollutant plays in jointly affecting ecosystem indicators, functions,
and services.  The clearest example  of this interaction is in assessment of the impacts of
acidifying deposition on aquatic ecosystems.
       Acidification in an aquatic ecosystem depends on the total acidifying potential of the
deposition of both N and S from both atmospheric deposition of oxides of nitrogen and sulfur as
well as the inputs from other sources of N and S such as reduced nitrogen and non-atmospheric
sources. It is the joint impact of the two pollutants that determines the ultimate effect on
organisms within the ecosystem, and critical ecosystem functions such as habitat provision and
biodiversity.  Standards that are set independently are less able to account for the contribution of
the other pollutant.  This suggests that interactions between oxides of nitrogen and oxides of
sulfur should be a critical element of the conceptual framework for ecologically relevant
standards. There are also important interactions between oxides of nitrogen and sulfur and
reduced forms of nitrogen, which also contribute to acidification and nutrient enrichment.
Although the standards do not directly address reduced forms of nitrogen in the atmosphere, e.g.
they do not require specific levels of reduced nitrogen, it is important that the structure of the
standards address the role of reduced nitrogen in determining the ecological effects resulting
from deposition of atmospheric oxides  of nitrogen and sulfur.  Consideration will also have to be
given to account for loadings coming from non-atmospheric sources  as ecosystems will respond
to these sources as well.
       In addition to the fundamental issues discussed above, the current structures of the
standards do not address the complexities in the responses of ecosystems to  deposition of oxides
of nitrogen and sulfur.  Ecosystems  contain complex groupings of organisms that respond in

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various ways to the alterations of soil and water that result from deposition of nitrogen and sulfur
compounds.  Different ecosystems therefore respond in different ways depending on a multitude
of factors that control how deposition is integrated into the system. For example, the same levels
of deposition falling on limestone dominated soils have a very different effect than those falling
on shallow glaciated soils underlain with granite. One system may over time display no obvious
detriment while the other may experience a catastrophic loss in fish communities.  This degree
of sensitivity is a function of many atmospheric factors which control rates of deposition as well
as ecological factors which control how an ecosystem responds to that deposition. The current
standards do not take into account spatial  and seasonal variations not only in depositional
loadings but also in sensitivity of ecosystems exposed to those loadings.
       The 2005 ambient conditions indicate that the current SC>2 and NC>2 secondary standards
are not exceeded (Figures 6-1 and 6-2) in  locations where ecological  effects have been observed,
and where critical loads of nitrogen and sulfur are exceeded.
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                          3-hr Max SO2 Concentrations (2005)
Figure 6-1.   Three hour average maximum 2005 SC>2 concentrations based on the SLAMS
reporting to EPA's Air Quality System (AQS) data base.   The current SC>2 secondary standard
based on the maximum 3 hour average value is 500 ppb, a value not exceeded.  While there are
obvious spatial gaps, the majority of these stations are located to capture maximum values
generally in proximity to major sources and high populations.  Lower relative values are
expected in more remote acid sensitive areas. See Table 2-1 for unit conversions
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                     Annual Average NO2 Concentrations (2005)
Figure 6-2    Annual average 2005 NC>2 concentrations based on the SLAMS reporting to
EPA's Air Quality System (AQS) data base.  The current NC>2 secondary standard is 53 ppb, a
value well above those observed. While there are obvious spatial gaps, the stations are located
in areas of relatively high concentrations in highly populated areas. Lower relative values are
expected in more remote acid sensitive areas.  See Table 2-1 for unit conversions
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6.3    ADVERSE EFFECTS ON THE PUBLIC WELFARE OCCURRING UNDER
       CURRENT AIR QUALITY CONDITIONS FOR NO2 AND SO2
       In the previous sections we have established that all areas of the U.S. were at
concentrations of 862 and NC>2 below the levels of the current standards. In many locations, 862
and NC>2 concentrations are substantially below the levels of the standards.  This pattern suggests
that levels of deposition and any effects on ecosystems due to deposition of oxides of nitrogen
and sulfur under recent conditions are occurring even though areas meet or are below current
standards. In this section we focus on summarizing the evidence of effects occurring at
deposition levels consistent with recent conditions.
       The ISA summarizes the available studies of relative nitrogen contribution and finds that
in much of the U.S., oxides of nitrogen contribute from 50 to 75 percent of total atmospheric
deposition relative to total reactive nitrogen that includes oxidized and reduced nitrogen species
(ISA Section 2.8.4). Although the proportion of total nitrogen loadings associated with
atmospheric deposition of nitrogen varies across locations (N deposition in the eastern U.S.
includes locations with greater than 9 kg N/ha-yr, and in the central U.S. high deposition
locations with values on the order of 6 to 7 kg N/ha-yr), the ISA indicates that atmospheric N
deposition is the main source of new anthropogenic N to most headwater streams, high elevation
lakes, and low-order streams. Atmospheric N deposition contributes to the total N load in
terrestrial, wetland, freshwater, and estuarine ecosystems that receive N through multiple
pathways. In several large estuarine systems, including the Chesapeake Bay, atmospheric
deposition accounts for between 10 and 40 percent of total nitrogen loadings (U.S. EPA, 2000).
       Atmospheric concentrations of oxides of sulfur account for nearly all S deposition in the
US. For the period 2004-2006, mean S deposition in the U.S. was greatest east of the
Mississippi River with the highest deposition amount, 21.3 kg S/ha-yr, in the Ohio River Valley
where most recording stations reported 3 year averages >10 kg S/ha-yr. Numerous other stations
in the East reported S deposition >5 kg S/ha-yr. Total S deposition in the U.S. west of the 100th
meridian was relatively low, with all recording stations reporting <2 kg S/ha-yr and many
reporting <1 kg S/ha-yr.  S was primarily deposited in the form of wet 864 2 followed in
decreasing order by a smaller proportion of dry 862 and a much smaller proportion of deposition
as dry SO42  .
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       New scientific evidence exists to address each of the areas of uncertainty raised in the
previous reviews.  Based on the new evidence, the current ISA concludes that:
          (1)    The evidence is sufficient to infer a causal relationship between acidifying
                 deposition (to which both oxides of nitrogen and sulfur contribute) and effects
                 on biogeochemistry related to terrestrial and aquatic ecosystems; and biota in
                 terrestrial and aquatic ecosystems.
          (2)    The evidence is sufficient to infer a causal relationship between N deposition,
                 to which oxidized and reduced nitrogen contribute, and the alteration of A)
                 biogeochemical cycling of N and carbon in terrestrial, wetland, freshwater
                 aquatic, and coastal marine ecosystems; B) biogenic flux of methane (CIHLi),
                 and N2O in terrestrial and wetland ecosystems; and C) species richness,
                 species composition, and biodiversity in terrestrial, wetland, freshwater
                 aquatic and coastal marine ecosystems.
          (3)    The evidence is sufficient to infer a causal relationship between S  deposition
                 and increased Hg methylation in wetlands and aquatic environments.
       Subsequent to the previous review of the secondary standard for oxides of nitrogen, a
great deal of information on the contribution of atmospheric deposition associated with ambient
oxides of nitrogen has become available. In Chapter 3 of the REA a thorough assessment is
provided of the contribution of oxidized nitrogen to nitrogen deposition throughout the U.S., and
the relative contributions of ambient oxidized and reduced forms of nitrogen. Staff concludes
that based on that analysis, ambient  oxides of nitrogen are a significant component of
atmospheric nitrogen deposition, even in areas with relatively high rates of deposition of reduced
nitrogen.  In addition, staff concludes that atmospheric deposition of oxidized nitrogen
contributes significantly to total nitrogen loadings in nitrogen sensitive ecosystems.
       As discussed throughout the REA document, there are several key  areas of risk that are
associated with ambient concentrations of oxides of nitrogen and sulfur. As noted earlier, in
previous reviews of the secondary standards for oxides of nitrogen and sulfur, the standards were
designed to protect against direct exposure of plants to ambient concentrations of the pollutants.
A significant shift in understanding of the effects of oxides of nitrogen and sulfur has occurred
since the last reviews, reflecting the large amount of research that has been conducted on the
effects of deposition of nitrogen and sulfur to ecosystems. The most significant risks  of adverse

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effects to public welfare are those related to deposition of oxides of nitrogen and sulfur to both
terrestrial and aquatic ecosystems.  These risks fall into two categories, acidification and nutrient
enrichment, which were emphasized in the REA are most relevant to evaluating the adequacy of
the existing standards in protecting public welfare from adverse ecological effects.

6.3.1   Acidification in sensitive aquatic and terrestrial ecosystems
       The focus of the REA case studies was on determining whether deposition of sulfur and
oxidized nitrogen in locations where ambient oxides of nitrogen and sulfur were at or below the
current standards was resulting in acidification and related effects. This review has focused on
identifying ecological indicators that can link atmospheric deposition to ecological effects
associated with acidification. Oxides of nitrogen and sulfur contribute to acidification in both
aquatic and terrestrial ecosystems,  although the indicators of effects differ.  Although there are
some geographic areas with both terrestrial and aquatic ecosystems that are vulnerable to
acidification, the case study areas do not fully overlap.   The locations of the case studies
evaluated in the REA are shown on Figure 6-3.
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                                           250  500   750  1,000
                                              Kilometers
Figure 6-3    National map highlighting the nine case study areas evaluated in the REA.
Aquatic Acidification
       Based on the case studies conducted for lakes in the Adirondacks and streams in
Shenandoah National Park, staff concludes that there is significant risk to acid sensitive aquatic
ecosystems at atmospheric concentrations of oxides of nitrogen and sulfur at or below the current
standards. This conclusion is based on application of the MAGIC model to estimate the effects
of deposition at levels consistent with atmospheric oxides of nitrogen and sulfur concentrations
that are at or below the current standards. An important ecological indicator for aquatic
acidification effects is acid neutralizing capacity (ANC) of a waterbody, and the case study
focused on evaluating whether locations were likely to be below critical values of ANC given
deposition levels associated with oxides of nitrogen and sulfur atmospheric concentrations that
meet the current standards.  In addition, the case studies assessed the ecological effects and some
of the known ecosystem services that are associated with different levels of ANC  in order to
associate levels of ANC with measures of public welfare that may be adversely affected by
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deposition levels consistent with atmospheric concentrations of oxides of nitrogen and sulfur that
meet the current standards.
       Staff concludes that the evidence and risk assessment support strongly a relationship
between atmospheric deposition of oxides of nitrogen and sulfur and loss of ANC in sensitive
ecosystems, and that ANC is an excellent indicator of aquatic acidification. Staff also concludes
that at levels of deposition associated with oxides of nitrogen and sulfur concentrations at or
below the current standards, ANC levels are expected to be below benchmark values that are
associated with significant losses in fish species richness (REA Section 4).
       Many locations in sensitive areas of the U.S. have ANC levels below benchmark levels
for ANC classified as severe, elevated, or moderate concern (see Figure 2-1). The average
current ANC  levels across 44 lakes in the Adirondack case study area is 62.1  jieq/L (moderate
concern). However, 44 percent of lakes had deposition levels exceeding the critical load for an
ANC  of 50 |ieq/L, and 28 percent of lakes had deposition levels exceeding the critical load for an
ANC  of 20 |ieq/L (REA Section 4.2.4.2).  This information indicates that almost half of the 44
lakes  in the Adirondacks case study area are at an elevated concern levels, and almost a third are
at a severe concern level. These levels are associated with greatly diminished fish species
diversity, and losses in the health and reproductive capacity of remaining populations. Based on
assessments of the relationship between number offish species and ANC level in both the
Adirondacks and Shenandoah areas, the number offish species is decreased by over half at an
ANC  level of 20 |ieq/L relative to an ANC level at 100 |ieq/L (REA Figure 4.2-1). At levels
below 20 |ieq/L, populations of sensitive species, such as brook trout, may decline significantly
during episodic acidification events.  When extrapolated to the full population of lakes in the
Adirondacks area using weights based on the EMAP probability survey (REA 4.2.6.1), 36
percent  of lakes exceeded the critical load for an ANC of 50  jieq/L and 13 percent of lakes
exceeded the  critical load for an ANC of 20 jieq/L.
       Many streams in the Shenandoah case study area also have levels of deposition that are
associated with ANC levels classified as severe, elevated, or moderate concern.  The average
ANC  under recent conditions for the 60 streams evaluated in the Shenandoah case study area is
57.9 |ieq/L, indicating moderate concern.  However,  85 percent of streams had recent deposition
exceeding the critical load for an ANC of  50 |ieq/L,  and 72 percent exceeded the critical load for
an ANC of 20 |ieq/L. As with the Adirondacks area, this information suggests that significant

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numbers of sensitive streams in the Shenandoah area are at risk of adverse impacts on fish
populations under recent conditions.  Many other streams in the Shenandoah area are likely to
experience conditions of elevated to severe concern based on the prevalence in the area of
bedrock geology associated with increased sensitivity to acidification suggesting that effects due
to stream acidification could be widespread in the Shenandoah area (REA 4.2.6.2).
       In the ISA it is noted that significant portions of the U.S. are acid sensitive, and that
current deposition levels exceed those that would allow recovery of the most acid sensitive lakes
in the Adirondacks (ISA Executive Summary). In addition, because of past loadings, areas of the
Shenandoah are sensitive to current deposition levels (ISA Executive Summary). Parts of the
West are naturally less sensitive to acidification and subjected to lower deposition (particularly
SOX) levels relative to the eastern United States, and as such, less focus in the ISA is placed on
the adequacy of the existing standards in these areas, with the exception of the mountainous
areas of the West, which experience episodic acidification due to deposition.
       While most (99 percent)  stream kilometers in the U.S. are not chronically acidified
under current conditions, a recent survey found sensitive streams in many locations in the U.S.,
including the Appalachian mountains, the Coastal Plain, and the Mountainous West (ISA Section
4.2.2.3).  In these sensitive areas, between 1 and 6 percent of stream kilometers are chronically
acidified.
       The ISA notes that "consideration of episodic acidification greatly increases the extent
and degree of estimated effects for acidifying deposition on surface waters." (ISA Section
3.2.1.6)  Some studies show that the number of lakes that could be classified as acid-impacted
based on episodic acidification is 2 to 3 times the number of lakes classified as acid-impacted
based on chronic ANC.  These episodic acidification events can have long term effects on fish
populations (ISA Section 3.2.1.6).  Under recent conditions, episodic acidification has been
observed in locations in the eastern U.S. and in the mountainous western U.S. (ISA Section
3.2.1.6).
       It can therefore be concluded that recent levels of oxides of nitrogen and sulfur are
associated with deposition that leads to ANC values below benchmark values known to cause
ecological harm in sensitive aquatic systems, including lakes and streams in multiple areas of the
U.S. These changes are known to have  impacts on ecosystem services including recreational
fishing which is discussed along with other services in Chapter 3. While other ecosystem

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services (e.g. habitat provisioning, subsistence fishing, and biological control as well as many
others) are potentially affected by reductions in ANC, confidence in the specific translation of
ANC values to these additional ecosystem services is much lower.

Terrestrial Acidification
       Based on the case studies on sugar maple and red spruce habitat, staff concludes that
there is significant risk to sensitive terrestrial ecosystems from acidification at atmospheric
concentrations of NOX and SOx at or below the current standards. This conclusion is based on
application of the simple mass balance model to deposition levels associated with NOX and SOx
concentrations at or below the current standards. The ecological indicator selected for terrestrial
acidification is the base cation to aluminum ratio (BC: Al), which has been linked to tree health
and growth. The results of the REA strongly support a relationship between atmospheric
deposition of NOX and SOx and BC: Al, and that BC:A1 is a good indicator of terrestrial
acidification. At levels of deposition associated with NOX and SOx concentrations at or below
the current standards,  BC:A1 levels are expected to be below benchmark values that are
associated with significant effects on tree health and growth. Such degradation of terrestrial
ecosystems could affect ecosystem services such as habitat provisioning, endangered species,
goods production (timber, syrup, etc.) and many others.
       Many locations in sensitive areas of the U.S. have BC:A1 levels below benchmark levels
classified as providing low to intermediate levels of protection to tree health. At a BC:A1 ratio of
1.2 (intermediate level of protection), red spruce growth can be reduced by 20 percent. At a
BC:A1 ratio of 0.6 (low level of protection), sugar  maple  growth can be decreased by 20 percent.
The REA did not evaluate broad  sensitive regions. However, in the sugar maple case study area
(Kane Experimental Forest), recent deposition levels are  associated with a BC: Al ratio below
1.2, indicating between intermediate and low level of protection, which would indicate the
potential for a greater than 20 percent reduction in growth.  In the red spruce case study area
(Hubbard Brook Experimental Forest), recent deposition  levels are associated with a BC: Al ratio
slightly above 1.2, indicating  slightly better than an intermediate level of protection (REA
Section 4.3.5.1).
       Over the full range of sugar maple, 12 percent of  evaluated forest plots exceeded the
critical loads for a BC:A1 ratio of 1.2, and 3 percent exceeded the critical load for a BC:A1 ratio

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of 0.6. However, there was large variability across states. In New Jersey, 67 percent of plots
exceeded the critical load for a BC:A1 ratio of 1.2, while in several states on the outskirts of the
range for sugar maple (e.g. Arkansas, Illinois) no plots exceeded the critical load for a BC:A1
ratio of 1.2.  For red spruce,  overall 5 percent of plots exceeded the critical load for a BC:A1 ratio
of 1.2, and 3 percent exceeded the critical load for a BC:A1 ratio of 0.6.  In the major red spruce
producing states (Maine, New Hampshire, and Vermont), critical loads for a BC:A1 ratio of 1.2
were exceeded in 0.5, 38, and 6 percent of plots.
       The ISA reported one study (McNulty, 1997) that estimated 15 percent of U.S. forest
ecosystems exceeded the critical loads for acidity for N and S deposition by >250 eq/ha/year
under current conditions (ISA Section 4.2.1.3). Staff concludes that this represents a significant
portion of sensitive terrestrial ecosystems.
       It can therefore be concluded that recent levels of NOX and SOx are associated with
deposition that leads to BC:A1 values below benchmark values that cause ecological harm in
some sensitive terrestrial ecosystems. While effects are more widespread for sugar maple, there
are locations with low to intermediate levels of protection from effects on both sugar maple and
red spruce.  While there are many other ecosystem services, including timber production, natural
habitat provision, and regulation of water, climate, and erosion, potentially affected by
reductions in BC: Al, linkages of BC: Al values to these additional ecosystem services is on the
whole not well understood.

6.3.2   Nutrient enrichment effects in sensitive aquatic and terrestrial ecosystems
       Nutrient enrichment effects are due to nitrogen loadings from both atmospheric and non-
atmospheric sources. Evaluation of nutrient enrichment effects requires an understanding that
nutrient inputs are essential to ecosystem health and that specific long term levels of nutrients in
a system affect the types of species that occur over long periods of time.  Short term additions of
nutrients can affect species competition, and even small additions of nitrogen in areas that are
traditionally nutrient poor can have significant impacts on productivity as well as species
composition.   Most ecosystems in the United States are nitrogen-limited, so regional decreases
in emissions and deposition of airborne nitrogen compounds could lead to some decrease in
growth of the vegetation that surrounds the targeted aquatic system but as discussed below
evidence for this is mixed. Whether these changes in plant growth are seen as beneficial or
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adverse will depend on the circumstances (e.g. managed ecosystems).  However, as noted earlier,
this review of the standards is focused on unmanaged ecosystems. Changes to unmanaged
systems may be adverse or beneficial depending on desired ecosystem services.  In assessing
adequacy of the current standards, we are focusing on the adverse effects of nutrient enrichment
in unmanaged ecosystems.  However, the following discussion provides a brief assessment of
effects in managed ecosystems.
       Nitrogen is a fundamental nutrient for primary production in both managed and
unmanaged ecosystems. Impacts of nutrient enrichment in managed ecosystems may be positive
or negative depending on the levels of nutrients from other sources in those areas.  Positive
growth effects can result when crops or commercial forests are not receiving enough nitrogen.
Nutrients deposited on crops from atmospheric sources are often referred to as passive
fertilization.  Most productive agricultural systems require external sources of nitrogen in order
to satisfy nutrient requirements.  Nitrogen uptake by crops varies, but typical requirements for
wheat and corn are approximately 150 kg/ha-yr and 300 kg/ha-yr, respectively (NAPAP,  1990).
Typical estimated rates of passive nitrogen fertilization are in the range of 0 to 5.5 kg/ha-yr
(NAPAP, 1991).
       Information on the effects of changes in passive nitrogen deposition on forestlands and
other terrestrial ecosystems is very limited. The multiplicity of factors affecting forests, including
other potential stressors such as ozone, and limiting factors such as moisture and other nutrients,
confound assessments of marginal changes in any one stressor or nutrient in forest ecosystems.
The ISA notes that only a fraction of the deposited nitrogen is taken up by the forests, most of
the nitrogen is retained in the soils (ISA 3.3.2.1). In addition, the ISA indicates that forest
management practices can significantly affect the nitrogen cycling within a forest ecosystem, and
as such, the response of managed forests to NOX deposition will be variable depending on the
forest management practices employed in a given forest ecosystem (ISA Annex C C.6.3)
Increases in the availability of nitrogen in N-limited forests via atmospheric deposition could
increase forest production over large non-managed areas, but the evidence is mixed, with some
studies showing increased production and other showing little effect on wood production (ISA
3.3.9). Because leaching of nitrate can promote cation losses, which in some cases create nutrient
imbalances, slower growth and lessened disease and freezing tolerances for forest trees, the net
effect of increased N on forests in the U.S. is uncertain (ISA 3.3.9).

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       In managed agricultural ecosystems, nitrogen inputs from atmospheric NOX comprise a
small fraction (less than 3 percent) of total nitrogen inputs, which include commercially applied
fertilizers as well as applications of composted manure.  And because of the temporal and spatial
variability in atmospheric deposition of NOX, it is unlikely that farmers would alter their
fertilization decisions based on expected nitrogen inputs from NOX.  And, in some locations,
farmers need less nitrogen inputs due to production of excess nitrogen through livestock. In
these locations, nitrogen production through livestock waste exceeds the absorptive capacity of
the surrounding land, and as such, excess nitrogen from deposition of NOX in those locations
reduces the capacity of the system to dispose of excess nitrogen, potentially increasing the costs
of waste management from livestock operations (Letson and Gollehon, 1996). A USD A
Economic Research Service report found that in 1997, 68 counties with high levels of confined
livestock production had manure nitrogen levels that exceed the assimilative capacity of the
entire county's crop and pasture land (Gollehon et al, 2001). In those locations, additional
nitrogen inputs from NOX deposition will result in excess nitrogen, leading to nitrogen leaching
and associated effects that adversely affect ecosystems.

Aquatic Nutrient Enrichment
       The REA case studies focused on coastal estuaries and revealed that while current
ambient loadings of atmospheric NOX are contributing to the overall depositional loading of
coastal estuaries, other non-atmospheric  sources are contributing in far greater amounts in total,
although atmospheric contributions are as large as some other individual source types. The
ability of current data and models to characterize the incremental adverse impacts of nitrogen
deposition is limited, both by the available ecological indicators, and by the inability to attribute
specific effects to atmospheric sources of nitrogen. The REA case studies used as the ecological
indicator for aquatic nutrient enrichment an index of eutrophication known as the Assessment of
Estuarine Trophic Status Eutrophi cation  Index (ASSETS El). This index is a  six level index
characterizing overall eutrophication risk in a waterbody. This indictor is not sensitive to
relatively large changes in nitrogen deposition. In addition, this type of indicator does not reflect
the impact of nitrogen deposition in conjunction with other sources of nitrogen.
       For example, if NOX deposition is contributing nine tenths of the nitrogen loading
required to move a waterbody from an ASSETS El category of "moderate" to a category of
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"poor", zeroing out NOX deposition will have no impact on the ASSETS El value. However, if
an area were to decide to put in place decreases in nitrogen loadings to move that waterbody
from "poor" to "moderate," the area would have to reduce the full amount of the loadings
through other sources if atmospheric deposition were not considered.  Thus, the adverse impact
of atmospheric nitrogen is in its contribution to the overall loading, and reductions in NOX will
decrease the amount of reductions from other sources of nitrogen loadings that would be required
to move from a lower ASSETS El category to a higher category. NOX deposition can also be
characterized as reducing the risk of a waterbody moving from a higher ASSETS El category to
a lower category, by reducing the vulnerability of that waterbody to increased loadings from
non-atmospheric sources.
       Based on the above considerations, staff preliminarily concludes that the ASSETS El is
not an appropriate ecological indicator for estuarine aquatic eutrophication. Staff further
concludes that additional analysis is required to develop an appropriate indicator for determining
the appropriate levels of protection  from N nutrient enrichment effects in estuaries related to
deposition of NOX. As a result, staff is unable to make a determination as to the adequacy of the
existing secondary NOX standard in protecting public welfare from N nutrient enrichment effects
in estuarine aquatic ecosystems.
       Additionally,  nitrogen deposition can alter species composition and cause eutrophication
in freshwater systems. In the Rocky Mountains, for example,  deposition loads of 1.5 to 2 kg/ha-
yr which are well within current ambient levels are known to cause changes in species
composition in diatom communities indicating impaired water quality (ISA Section 3.3.5.3).  It
then seems apparent then that the existing secondary standard  for NOX does not protect such
ecosystems and their resulting services from impairment.

Terrestrial Nutrient Enrichment
       The scientific literature has many examples of the deleterious effects caused by excessive
nitrogen loadings to terrestrial systems. Several studies have set benchmark values for levels of
N deposition at which scientifically adverse effects are known to occur.  These benchmarks are
discussed  more thoroughly in Chapter 5 of the REA.  Large areas of the  country appear to be
experiencing deposition above these benchmarks for example, Fenn et al. (2008) found that at
3.1 kg N/ha-yr, the community of lichens begins to change from acidophytic to tolerant species;
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at 5.2 kg N/ha-yr, the typical dominance by acidophytic species no longer occurs; and at 10.2 kg
N/ha-yr, acidophytic lichens are totally lost from the community. Additional studies in the
Colorado Front Range of the Rocky Mountain National Park support these findings and are
summarized in Chapter 6.0 of the Risk and Exposure Assessment. These three values (3.1, 5.2,
and 10.2 kg/ha-yr) are one set of ecologically meaningful benchmarks for the mixed conifer
forest (MCF) of the pacific coast regions. Nearly all of the known sensitive communities receive
total nitrogen deposition levels above the 3.1 N kg/ha-yr ecological benchmark according to
the 12 km, 2002 CMAQ/NADP data, with the exception of the easternmost Sierra Nevadas.
MCFs in the southern portion of the Sierra Nevada forests and nearly all MCF communities in
the San Bernardino forests receive total nitrogen deposition levels above the 5.2 N kg/ha-yr
ecological benchmark.
       Coastal Sage Scrub communities (CSS) are also known to be sensitive to community
shifts caused by excess nitrogen loadings. Wood et al. (2006) investigated the amount of
nitrogen utilized by healthy and degraded CSS  systems. In healthy stands, the authors estimated
that 3.3 kg N/ha-yr was used for CSS plant growth (Wood et al., 2006). It is assumed that 3.3 kg
N/ha-yr is near the point where nitrogen is no longer limiting  in the CSS community. Therefore,
this amount can be considered an ecological benchmark for the CSS community. The majority of
the known CSS range is currently receiving deposition in excess of this benchmark. Thus, staff
concludes that recent conditions where NOX ambient concentrations are at or below the current
NOX secondary standards are not adequate to protect against anticipated adverse impacts from N
nutrient enrichment in sensitive ecosystems.

6.3.3   Other Ecological Effects (Eg. Mercury Methylation) Associated With Deposition of
       Atmospheric Oxides Of Nitrogen and/or Sulfur
       It is stated in the ISA (ISA Sections 3.4.1 and 4.5) that mercury is a highly neurotoxic
contaminant that enters the food web as a methylated compound, methylmercury. Mercury is
principally methylated by sulfur-reducing bacteria and can be taken up by microorganisms,
zooplankton and macroinvertebrates. The contaminant is concentrated in higher trophic levels,
including fish eaten by humans. Experimental evidence has established that only inconsequential
amounts of methylmercury can be produced in  the absence of sulfate. Once methylmercury is
present, other variables influence  how much accumulates in fish, but elevated mercury levels in
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fish can only occur where substantial amounts of methylmercury are present. Current evidence
indicates that in watersheds where mercury is present, increased oxides of sulfur deposition very
likely results in additional production of methylmercury which leads to greater accumulation of
MeHg concentrations in fish (Munthe et al, 2007; Drevnick et al, 2007).
       The production of meaningful amounts of methylmercury (MeHg) requires the presence
of SC>42" and mercury, and where mercury is present, increased availability of SC>42" results in
increased production of MeHg. There is increasing evidence on the relationship between sulfur
deposition and increased methylation of mercury in aquatic environments; this effect occurs only
where other factors are present at levels within a range to allow methylation. The production of
methylmercury requires the presence of sulfate and mercury, but the amount of methylmercury
produced varies with oxygen content, temperature, pH, and supply of labile organic carbon (ISA
Section 3.4). In watersheds where changes in sulfate deposition did not produce an effect, one or
several of those interacting factors were not in the range required for meaningful methylation to
occur (ISA Section 3.4). Watersheds with conditions known to be conducive to mercury
methylation can be found in the northeastern United States and southeastern Canada. The
relationship between sulfur and methylmercury production is addressed qualitatively in Chapter
6 of the Risk and Exposure Assessment.
       With respect to sulfur deposition and mercury methylation, the final ISA determined: The
evidence is sufficient to infer a causal relationship between sulfur deposition and increased
mercury methylation in wetlands and aquatic environments.  However, staff did not conduct a
quantitative assessment of the risks associated with increased mercury methylation under current
conditions. As such, staff is unable to make a determination as to the adequacy of the existing
SC>2 standards in protecting  against welfare effects associated with increased mercury
methylation.
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6.4    REFERENCES:

Drevnick, P.E., D.E. Canfield, P.R. Gorski, A.L.C. Shinneman, D.R. Engstrom, D.C.G. Muir,
       G.R. Smith, PJ. Garrison, L.B. Cleckner, J.P. Hurley, R.B. Noble, R.R. Otter, and J.T.
       Oris. 2007. Deposition and cycling of sulfur controls mercury accumulation in Isle
       Royale fish. Environmental Science and Technology ¥7(21):7266-7272.
Fenn, M.E., S. Jovan, F. Yuan, L. Geiser, T. Meixner, and B.S. Gimeno. 2008. Empirical and
       simulated critical loads for nitrogen deposition in California mixed conifer forests.
       Environmental Pollution 755(3):492-511.
Munthe, 1, R.A. Bodaly, B.A. Branfireun, C.T. Driscoll, C.C. Gilmour, R. Harris, M. Horvat, M.
       Lucotte, and O. Malm. 2007. Recovery of mercury-contaminated fisheries. Ambio 36:33-
       44.
U.S. EPA (Environmental Protection Agency). 1982. Review of the National Ambient Air Quality
       Standards for Sulfur Oxides: Assessment of Scientific and Technical Information.
       OAQPS Staff Paper. EPA-450/5-82-007.  U.S. Environmental Protection Agency, Office
       of Air Quality Planning and Standards, Research Triangle Park, NC.
U.S. EPA (Environmental Protection Agency). 1995. Review of the National Ambient Air Quality
       Standards for Nitrogen Dioxide: Assessment of Scientific and Technical Information.
       OAQPS Staff Paper. EPA-452/R-95-005. U.S. Environmental Protection Agency, Office
       of Air Quality Planning and Standards, Research Triangle Park, NC. September.
U.S. EPA. (Environmental Protection Agency). 2000. Deposition of Air Pollutants to the Great
       Waters: Third report to Congress. (EPA-453/R-00-005)
       http://www.epa.gov/air/oaqps/gr8water/3rdrpt/index.html; Research Triangle Park, NC;
       Office of Air Quality Planning and Standards;
       U.S. Environmental Protection Agency
U.S. EPA (Environmental Protection Agency). 2008. Integrated Science Assessment (ISA) for
       Oxides of Nitrogen and Sulfur-Ecological Criteria (Final Report). EPA/600/R-
       08/082F. U.S. Environmental Protection Agency, National Center for Environmental
       Assessment-RTF Division, Office of Research and Development, Research Triangle
       Park, NC. Available at http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=201485.
Wood, Y., T. Meixner, P. J. Shouse, and E.B. Allen. 2006. Altered Ecohydrologic response
       drives native shrub loss under conditions of elevated N-deposition. Journal of
       Environmental Quality 35:76-92.
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         7      CONSIDERATION OF ALTERNATIVE STANDARDS
                         FOR AQUATIC ACIDIFICATION
       Having reached the conclusion that the current NC>2 and SC>2 standards are not adequate
to provide appropriate protection against deposition-related effects associated with oxides of
nitrogen and sulfur, staff considers alternative standards that are multi-pollutant and multimedia
in nature to address such effects on public welfare.  The inherently complex and variable
linkages between ambient concentrations of nitrogen and sulfur oxides, the related deposited
forms of nitrogen and sulfur, and the ecological responses that are associated with public welfare
effects call for consideration of an ecologically relevant design of a standard that reflects these
linkages.  Such a standard will necessarily be more complex than the NAAQS that have been set
historically to address effects associated with direct exposure to a single pollutant. Nonetheless,
an ecologically relevant multi-pollutant, multimedia standard to address deposition-related
effects is still appropriately defined in terms of the same basic elements that are used to define
any NAAQS - indicator, form, averaging time, and level - with the form incorporating
additional structural elements that reflect these multi-pollutant and multimedia attributes.  These
structural elements include the use of an ecological indicator, tied to the ecological effect we are
focused on, that accounts for ecologically relevant factors other than ambient  air concentrations.
All of these elements are needed to enable a linkage from ecological indicator to ambient air
indicators to completely define an ecologically relevant standard.
       More specifically, in this chapter we focus on the development of an ecologically
relevant standard to address effects associated with acidifying deposition of oxides of nitrogen
and sulfur in sensitive aquatic ecosystems. This focus is consistent with the information
presented in the ISA and RE A which highlighted the sufficiency of the quantity and quality of
the available evidence and assessments associated with aquatic acidification relative to the
information and assessments available for other deposition-related effects, including terrestrial
acidification and aquatic and terrestrial nutrient enrichment. Based on its review of these
documents, CAS AC agreed with the conclusion that aquatic acidification should be the focus for
developing a new multi-pollutant, multimedia standard in this review. In reaching conclusions
about a standard designed to address aquatic acidification effects, we also recognize that such a
standard may also provide some degree of protection against other deposition-related effects,
drawing from information presented above in chapter 5.

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        Our development of an alternative standard for aquatic acidification recognizes the need
for a nationally applicable standard for protection against adverse effects of aquatic acidification
to public welfare, while recognizing the complex and heterogeneous interactions between
ambient air concentrations of nitrogen and sulfur oxides, the related deposition of nitrogen and
sulfur, and associated ecological responses. Our approach also recognizes that while a standard
is national in scope and coverage, the effects to public welfare from aquatic acidification will not
occur to the same extent in all locations in the U.S., given the inherent variability of aquatic
systems to the effects of acidification.  As noted above in chapters 3  and 4, many locations in the
U.S. are naturally protected against acid deposition due to underlying geological conditions.
Likewise, some locations in the U.S., including lands managed for commercial agriculture and
forestry, are not likely to be negatively impacted by current  levels of nitrogen and sulfur
deposition.  As a result, while the alternative standard we are considering would apply
everywhere, it is structured to account for differences in the sensitivity of ecosystems across the
country.  This  allows for appropriate protection of sensitive  aquatic ecosystems, which are
relatively pristine and wild and generally  in rural areas, and  the services provided by such
sensitive ecosystems,  without requiring more protection than is needed elsewhere.
       In this  chapter we present our reasoning for developing a standard that employs (1) NOy
and SOx as the atmospheric indicators; (2) a form that takes into account variable factors, such
as atmospheric and ecosystem conditions that modify the amounts of deposited oxides of
nitrogen and sulfur; the distinction between oxidized and reduced forms of nitrogen;  effects of
deposited nitrogen and sulfur on aquatic ecosystems in terms of the ecological indicator ANC;
and the representativeness of water bodies within a defined  spatial area; (3) a multi-year
averaging time, and (4) a standard level defined in terms of  a target ANC level that, in the
context of the  above form, identifies the allowable levels of concentrations of NOy and SOx in
the ambient air. In developing such a standard, we have  defined an index, termed an aquatic
acidification index (AAI), directly expressed in terms of atmospheric concentrations  of oxides of
nitrogen and sulfur, that can be applied across the country to convey the allowable levels of
oxides of nitrogen and sulfur in the ambient air based on various factors such as the sensitivity of
an area and the desired degree of protection from aquatic acidification caused by atmospheric
deposition.
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       In considering such an alternative standard for aquatic acidification, we focus on each
element of the standard, including alternative indicators (section 7.1), forms (section 7.2),
averaging times (section 7.3), and levels (section 7.4). We then consider implications of various
combinations of alternative forms and levels, by characterizing areas that currently would likely
not meet such standards (section 7.5). A summary of staff conclusions with regard to the current
standards and alternative standards that are appropriate for consideration in this review is
presented in section 7.6.

7.1    INDICATORS
       In considering alternative ambient air indicators, we primarily focus on the important
attribute of association. Association in a broad sense refers to how well an ambient air indicator
relates to the ecological effects of interest by virtue of the conceptual or process-based
framework, linking indicator and effects as well as through empirical evidence.  We also
consider how measurable or quantifiable an indicator is to enable its use as an effective  indicator
of relevant ambient air concentrations.
       As discussed above in chapter 6, staff concludes that indicators other than NC>2 and SC>2
should be considered as the appropriate indicators of oxides of nitrogen and sulfur in the ambient
air for protection against the acidification effects associated with deposition of oxides of nitrogen
and sulfur.  This conclusion  is based on the recognition that all forms of nitrogen and sulfur
in the ambient air contribute to deposition and resulting acidification,  and as such,  N02
and S02 are incomplete indicators In principle,  the indicators should represent the
species associated with oxides of nitrogen  and sulfur in the ambient air that can contribute
acidifying deposition.  This includes both the species of oxides of nitrogen and sulfur that
are directly emitted as well  as species transformed in the atmosphere that retain the
nitrogen and sulfur atomsfrom directly emitted oxides of nitrogen and sulfur.   We
emphasize the individual atoms associated with oxides of nitrogen and sulfur because the
acidifying potential is specific to nitrogen and sulfur, and not other atoms (e.g., H, C,  0)
whether derived from the original source of oxides of  nitrogen and sulfur emissions or from
atmospheric transformations  For  example,  the acidifying potential of each molecule of
N02, NO,  HN03orPAN  is identical, as is the potential  for each molecule of S02 or ion
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of participate sulfate,  S04.  Each atom of sulfur affords twice the acidifying potential of
any atom of nitrogen.
       The next two sub-sections address indicators for oxides of sulfur and nitrogen,
respectively.  The discussion on sulfur is brief,  reflecting a general lack of issues or
alternatives for this pollutant, while the discussion on oxides of nitrogen addresses several
related issues and alternatives
7.1.1   Ambient air indicators for oxides of sulfur
       Oxides of sulfur include the gases sulfur monoxide (SO), sulfur dioxide (S02),
sulfur trioxide (S03),  disulfur monoxide (820), and particulate-phase sulfur compounds
that result from gas-phase sulfur oxides interacting with particles  However, the sum of SO2
and SC>4 does represent virtually the entire ambient air mass of sulfur that contributes to
acidification. In addition to accounting for virtually all the potential for acidification from
oxidized sulfur in the ambient air, there are reliable methods to monitor the concentrations of
SO2 and particulate 864. In addition, much of the data used to develop the technical basis for the
standard is based on monitoring or modeling of these species.l  Staff concludes that the sum of
SC>2 and SC>4, referred to as SOx, are appropriate ambient air indicators of oxides of sulfur
because they represent virtually all of the acidification potential  of ambient air oxides of sulfur
and there are reliable methods suitable for measuring SO2 and SC>4.
7.1.2   Ambient air indicators for oxides of nitrogen
       Total reactive oxidized nitrogen, NOy, defined in chapter 2,  incorporates basically
all of the oxidized nitrogen species that have acidifying potential and as such, NOy should
be considered as an appropriate indicator for oxides of nitrogen.  NOy is an aggregate
measure of NOx (defined as NO and NO2) and all of the reactive oxidized products of NOx.
That is, NOy is a group of nitrogen compounds in which all of the compounds are either an oxide
of nitrogen or the nitrogen atoms in the compounds that came from oxides of nitrogen.  NOy is
especially relevant as an acidification indictor in that it both relates to the oxides of nitrogen in
1 As discussed above in chapter 2, we note that SO2 and paniculate SO4 are routinely measured in ambient air
monitoring networks, although only CASTNET filter packs do not intentionally exclude particle size fractions. The
CMAQ treatment of SOx is the simple addition of both species, which are treated explicitly in the model
formulation.  All particle size fractions are included in the CMAQ SOx estimates.
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the ambient air and also represents the acidification potential of all oxidized nitrogen species in
the ambient air, whether an oxide of nitrogen or derived from oxides of nitrogen.
       There are currently available reliable methods of measuring NOy.  The term "aggregate"
measure means that the NOy as measured is not based on measuring each individual species of
NOy and calculating an NOy value by addition (as performed in CMAQ processing), but rather
produces a measurement, as described in chapter 2, in which all of the individual NOy species
are processed by the measurement technique to produce a single aggregate measure of all of the
nitrogen atoms that were associated with any NOy species.  Consequently, the NOy
measurement effectively provides the sum of all individual species, but the identity of the
individual species is lost. As discussed above, the accounting for the individual nitrogen atoms
is an accounting of the ambient air acidification potential of oxides of nitrogen and their
transformation products and therefore the most relevant indicator for aquatic acidification effects
associated with oxides of nitrogen.
       However, the loss of the information on individual species has motivated consideration of
alternative  or more narrowly defined indicators for oxides of nitrogen.  Considering a subset of
NOy species is based on the following lines of reasoning.  First, the actual dry deposition of
nitrogen is  determined on an individual species basis by multiplying the species concentration
times a species-specific deposition velocity and then summed to develop an estimate of total dry
deposition.  Consequently, the use of individual ambient species has the potential to be more
consistent with the underlying  science of deposition and, therefore, has the potential to allow for
a more rigorous evaluation of dry deposition with specialized field studies. In addition, there has
been a suggestion of focusing only on the most quickly depositing NOy species, such as nitric
acid and HNOs, as contributions from other NOy species such as NO2 may be negligible. These
alternative  indicators are considered below.
•   What are  the relative merits of using the sum of each individual NOy species as the
    indicator for oxides of nitrogen?
       Dry deposition of NOy is treated as the sum of the deposition of each individual species
in advanced process-based air quality models like CMAQ, as described in chapter 2.
Conceptually one could extend this process-based approach by using all NOy species
individually as separate indicators for oxides of nitrogen and requiring, for example,
measurements of each of the species, including the dominant species of HNOs, particulate

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nitrate, true NC>2, NO, and PAN.  The potential attraction of using individual species would be
the reliance on actual deposition velocities that have more physical meaning in comparison to a
model constructed aggregate deposition of NOy, which is difficult to evaluate with observations
because of the assimilation of many species with disparate deposition behavior. The major
drawback of using individual NOy species as the indicators is the  lack of reliable measurement
techniques, especially for PAN and NO2 in rural locations, which  renders the use of virtually any
individual NOy species, except for NO and perhaps particulate nitrate, as functionally inadequate
from a measurement perspective.
•  What are the relative merits of using a subset of NOy species as the indicators for
   oxides of nitrogen?
       If certain species provide relatively minor contributions to total NOy deposition, then we
could consider excluding them as part of the indicator.   As discussed in chapter 2, each nitrogen
species within the array of NOy species has species-specific dry deposition velocities.  For
example, the deposition velocity of nitric acid is much greater than the velocity for nitrogen
dioxide and, consequently,  for a  similar ambient air concentration, nitric acid contributes more
deposition of acidifying nitrogen relative to nitrogen dioxide.  In transitioning from source-
oriented urban locations to  rural  environments, the ratio of the concentrations of nitric acid and
PAN to nitrogen dioxide increases.
       Based on the reasoning that a larger fraction of the deposited NOy is accounted for by
total nitrate (the sum of nitric acid and particulate nitrate), a surrogate for more rapidly
depositing fraction of NOy, combined with the availability of reliable total nitrate measurements
through CASTNET, consideration  has been given to using total nitrate as the indicator for oxides
of nitrogen.  The use of total nitrate as it could be incorporated within the form of the standard is
described in appendix E. One can reason that nitrate correlates well with total reactive oxidized
nitrogen deposition relative to NOy (as discussed in chapter 2), given the inherent noise
associated with variable contributions of low deposition velocity species (e.g., NO2) that may
have relatively high ambient concentrations.  However, modeling  simulations  suggest that NOy
may be a more robust indicator in terms of absolute changes in ambient air concentrations,
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relative to nitric acid2, of relating changes in an ambient air indicator to changes in nitrogen
deposition driven by changes in ambient concentrations of oxides of nitrogen (Figure 2-32).
       In summary, the disadvantages of using a subset of NOy species such as total nitrate or
nitric acid as an indicator are that a significant portion of ambient mass with the potential for
acidifying deposition is not captured.  This conclusion relates to species with high and low
deposition velocities. For example, nitrogen dioxide (low deposition velocity) alone is an
inadequate indicator by itself because it generally corresponds to less than 50% of the ambient
air contribution to nitrogen deposition, yet is a necessary piece of the aggregated total NOy
indicator because it can contribute significantly to deposited nitrogen.
•  What are staff conclusions with regard to an appropriate indicator for oxides of
   nitrogen?
       Staff concludes that total reactive oxidized nitrogen, NOy, is the appropriate ambient
indicator based on its direct relationship to deposition associated with aquatic acidification and
its direct relationship to oxides of nitrogen in the ambient air.  Because NOy represents all of the
potentially acidifying oxidized nitrogen species in the ambient air, it is appropriately associated
with the deposition of potentially acidifying oxides of nitrogen. In addition, there are reliable
methods available to measure NOy.  Measurement of each individual species of NOy, or the
measurement of only a subset of species of NOy is less appropriate, because a subset would fail
to account for significant portions of the oxidized reactive nitrogen that relates to acidification,
and because there are not reliable measurements methods available to measure all of the
individual species of NOy.
       NOy also is a useful measurement for model evaluation purposes.  Model evaluation
considerations are especially important, recognizing the unique role that CMAQ provides as
described below in section 7.2. Both of these data uses, NAAQS indicator and model evaluation,
are best served through characterizing the total mass of all relevant species, which in the case of
oxidized nitrogen, is best reflected through NOy.
       In reaching this conclusion, we note that the use of NOy, as well as SO2 and SO/t, as the
ambient air indicators for a new aquatic acidification standard for oxides of nitrogen and sulfur
2Nitric acid concentrations are significantly higher than paniculate nitrate and served as a more convenient
regression variable without affecting the correlations.
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was introduced in the first draft of this PA and was generally supported by CAS AC in its review
of that document (Russell and Samet, 2010a).
7.1.3   Monitoring Considerations
       Monitoring considerations specific to the indicators discussed above are briefly
summarized here, while other monitoring related issues are addressed in more detail above in
chapter 2. With respect to the oxides of sulfur indicators, 862 and 864, and oxides of nitrogen
indicator, NOy, we recognize that reliable monitoring methods are currently available.  Protocols
for monitor operations and quality assurance would be provided in conjunction with the
establishment of a standard that utilized these ambient indicators for oxides of nitrogen and
sulfur.
       With regard to the design of appropriate monitors for particulate sulfate, staff concludes
that an instrument design that does not intentionally omit any size fraction of the entire range of
sulfate particle diameters is appropriate to consider for measuring sulfate.  This inclusion of all
particle diameters considers that while most of the potential depositing mass is largely
incorporated in those particles with aerodynamic diameters less than 2.5 |i, particles with larger
diameters deposit much more quickly and can be an important contributor to sulfate deposition in
certain areas.
       Further, with regard to monitoring instrumentation for SO2, staff concludes that available
SC>2 continuously operating instruments with the ability to capture low concentrations  such as
those deployed in NCore are appropriate to consider.  In addition, consideration should be given
to the CASTNET filter pack (FP) 862 measurement technique for the purpose of this secondary
standard, since the 1-week averaging period employed by the CASTNET sampler is adequate for
the purposes of a standard based on annual average calculations (as discussed below in section
7.3).

7.2     FORM
       Based on the evidence of the aquatic acidification effects caused by NOy and SOx, staff
concludes that it is appropriate to develop a new form that is ecologically relevant for addressing
deposition-related effects, specifically for aquatic acidification effects.  EPA staff has developed
a conceptual design for the form of the standard that includes three main components:  ecological
indicators, deposition metrics that relate to ecological indicators, and a function that relates
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ambient indicators to deposition metrics.  Collectively, these three components link the
ecological indicator to ambient indicators, as illustrated in Fig 7-1.
                         Linking atmospheric
                         oxides of S and  N deposition to
                         ecological indicator
  Ecological effects and                                    Linking deposition to "allowable"
  ecological indicator                                      concentrations of ambient air
                                                        indicators of oxides of N and S

Figure 7-1.   Conceptual design of the form of an aquatic acidification standard for oxides
of nitrogen and sulfur.

       The simplified flow diagram in Figure 7-1 compresses the various atmospheric,
biological, and geochemical processes associated with acidifying deposition to aquatic
ecosystems (Figure 2-1) into a simplified conceptual picture. The ecological indicator (left box)
is related to atmospheric deposition through biogeochemical ecosystem models (middle box),
which associate a target deposition load to a target ecological indicator.  Once a target
deposition is established, associated allowable air concentrations are determined (right box)
through the relationships between concentration and deposition that are embodied in air quality
models such as CMAQ.  In sections 7.2.1 - 7.2.3 we describe the development and rationale for
each of these components.   Section 7.2.4 ties together these three components and develops the
full expression of the form of the standard using the concept of a national "aquatic acidification
index" that represents a target ANC level as a function of ambient air concentrations.  Section
7.2.5  addresses spatial aggregation issues associated with defining each of the terms of this index
within a spatial area.
        The aquatic acidification index (AAI)3 is designed to be an ecologically relevant form of
the standard that determines the allowable levels of ambient NOy and SOx based on a target
ANC limit for the U.S. The intent of the AAI is to weight atmospheric concentrations of oxides
of nitrogen and sulfur by their propensity to contribute to acidification through deposition, given
the fundamental  acidifying potential of each pollutant, and to take into account the ecological
factors that govern acid sensitivity in different ecosystems.  The index also accounts for the
contribution of reduced nitrogen to acidification. Thus, the AAI encompasses those attributes of
3 This index was previously termed the aquatic acidification protection index (AAPI) in earlier documents.
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specific relevance to protecting ecosystems from the acidifying potential of ambient air
concentrations of NOy and SOx.
       Staff notes two important concepts illustrated by the Os and PMio NAAQS that lend
support to using this index as the form of a NAAQS.  First, in recent reviews of the secondary
ozone standards, EPA has considered use of a form of the standard that reflects ecologically
relevant exposures,  by using a  cumulative index which weights exposures at higher
concentrations greater than those at lower concentrations based on scientific literature
demonstrating the cumulative nature of Os-induced plant effects and the need to give greater
weight to higher concentrations (EPA, 2007; See 75 FR 2938, 2999 January 19, 2010). Staff
also notes that PMio is the indicator for the coarse PM NAAQS standard (PMio = PM2.5 +PM i0.
2.5). Although the standard has a single level (150 ug/m3), the actual amount of coarse PM that is
allowed varies depending on how much fine PM (PM2.5) is present. By its nature, the PMio
NAAQS provides the appropriate protection from exposure to coarse PM across locations using
PMio as the indicator, by allowing the level of coarse PM to vary appropriately across locations.
The form for a standard for oxides of nitrogen and sulfur developed in this assessment builds on
this concept by using an index that weights the ambient concentrations to reflect their
relationship to acidification and provides the appropriate protection across the country by
allowing ambient air concentrations of oxides of nitrogen and sulfur to vary appropriately based
on ecosystem sensitivity and other relevant factors.
7.2.1   Ecological indicator
          Ecological effects and
          ecological indicator
       In considering alternative ecological indicators, we again primarily focus on the
important attribute of association.  In the case of an ecological indicator for aquatic
acidification, association refers to the relationship between the indicator and adverse effects as
discussed in chapter 3. Because of the conceptual structure of the form of this standard (Figure
7-1), this particular ecological indicator must also link up in a meaningful and quantifiable
manner with acidifying atmospheric deposition. In effect, the ecological indicator for aquatic
acidification is the bridge between the biological impairment we are focused on and deposition
of NOy and SOx.
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       This section presents the staffs rationale for selecting acid neutralizing capacity (ANC)
as the appropriate ecological indicator for consideration. Recognizing that ANC is not itself the
causative or toxic agent for adverse aquatic acidification effects, the rationale for using ANC as
the relevant ecological indicator is based on the following:
   •   ANC is directly associated with the causative agents, pH and dissolved aluminum, both
       through empirical evidence and mechanistic relationships;
   •   Empirical evidence shows very clear and strong relationships between adverse effects
       and ANC;
   •   ANC is a more reliable indicator from a modeling perspective, allowing use of a body of
       studies and technical analyses related to ANC and acidification to inform the
       development of the standard; and
   •   ANC literally embodies the concept of acidification as posed by the basic principles of
       acid base chemistry and the measurement method used to estimate ANC and, therefore,
       serves as a direct index to  protect against acidification.

•  What ecological indicators are appropriate to relate the effects of acidifying deposition
   to aquatic systems?
       Ecological indicators of acidification in aquatic ecosystems can be chemical or biological
components of the ecosystem that are altered by the acidifying effects of nitrogen and sulfur
deposition. A desirable ecological indicator for aquatic acidification is one that is measurable or
estimable, linked causally to deposition of nitrogen and sulfur, and linked  causally, either
directly or indirectly to ecological effects known or anticipated to adversely affect public
welfare.
       As summarized in chapter 2, atmospheric deposition of NOy and SOx causes aquatic
acidification through the input of  strong acid anions (e.g. N(V and SC>42")  that ultimately shifts
the water chemistry equilibrium toward increased hydrogen ion levels (or  decreased pH).   The
anions are deposited either directly to the aquatic ecosystem, or indirectly  via transformation
through soil nitrification processes and subsequent  drainage from terrestrial ecosystems. In other
words, when these anions are mobilized in the terrestrial soil, they can leach into adjacent water
bodies.  Aquatic acidification is indicated by changes in the surface water  chemistry of
ecosystems. In turn, the alteration of surface water chemistry has been linked to negative effects
on the biotic integrity of freshwater ecosystems. There is a suite of chemical indicators that could
be used to assess the effects of acidifying deposition on lake or stream acid-base chemistry.
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These indicators include acid neutralizing capacity (ANC); alkalinity (ALK); base neutralizing
capacity, commonly referred to as acidity (ACY); surface water pH; trivalent aluminum, Al+3;
concentrations of major anions (SC>42", N(V);  cations (Ca2+, Mg+2, K+); or sums of cations or
anions.
       Indictors such as specific anions, cations, or their groupings, while relevant to
acidification processes, are not robust acidification indicators as it is the relative balance of
cations and anions that is more directly associated with acidification.  That balance is captured
by ANC and ALK.  Acidity, ACY, is the corollary of ANC from the perspective of how much
strong base it takes to reach an equivalence point. Because ACY is not used in most ecosystem
assessments, the body  of information relating ACY to effects is too limited to serve as a basis for
an appropriate ecological indicator. Aluminum and other metals are causative toxic agents that
directly impair biological functions.  However, aluminum, or metals in general, have high
variability in concentrations that can be linked to effects, often at extremely low levels which in
some cases approach detectability limits, exhibit rapid transient responses, and are often
confounded by the presence of other toxic metals. These concerns limit the use of metals as
reliable and measurable ecological indicators.  Hydrogen ion (H+) concentrations, using their
negative logarithmic values, or pH, are well correlated with adverse effects, as discussed in
chapter 3, and determine the solubility of metals such as aluminum (figure 3-1). However, pH
also is not a preferred acidification indicator due to its highly transient nature and other concerns,
which are discussed below in the  context of why ANC is a preferred indictor.
       ANC and ALK are very similar quantities and are used interchangeably in the literature
and for some of the analyses presented in this document. Both ANC and ALK are defined as the
amount of strong acid required to reach a specified equivalence point (Stumm and  Morgan,
1981).  For acid base solutions, an equivalence point can be thought of as the point to which the
addition of strong acids (i.e., titration) is no  longer neutralized by the solution4.  This explains
the term acid neutralizing capacity, or ANC, as ANC truly relates directly to the capacity of a
system to neutralize acids.  The differences between ANC and ALK are based on operational
definitions and subject to various interpretations as described by Hemond, 1992.  ANC  is a
4 The common equivalent point that defines ANC or ALK is based on a solution of water, absent dissolved organic
carbon (DOC), that is open to the atmosphere for CO2 exchange and dissolved inorganic carbon (DIG) exists in the
form of bicarbonate ion, HCO3", carbonate ion, CO3"2 and carbonic acid, H2CO3. In such a solution, when the
titration with strong acid brings the pH down to the equivalence point at a pH of 4.5, all inorganic carbon is fully
protonated as carbonic acid, H2CO3.
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preferred over ALK as the body of scientific evidence has focused on ANC and effects
relationships.  ALK is more widely associated with more general characterizations of water
quality such as the relative hardness of water associated with carbonates.
       Having reasoned that ANC is a preferred indicator to ALK, ACY, individual metals or
groupings of ions, we consider the relative merits of ANC compared to pH, which is a well
recognized indicator of acidity and a more direct causative agent with regard to adverse effects.
First, the linkage between ANC  and pH is considered in recognition of the causative association
between pH and effects.
•  What is the mechanistic basis for and empirical evidence of the association between pH
   and ANC?
       ANC directly is not the causative toxic agent impacting aquatic species diversity.   The
scientific literature generally emphasizes the links between pH and adverse effects as described
in chapter 3.  It is important, therefore, to establish that ANC and pH are well related from a
mechanistic perspective as well through empirical evidence.   ANC and pH are co-dependent on
each other based on the requirement that all solutions are electrically neutral, meaning that any
solution must satisfy the condition that all negatively charged species must be balanced by all
positively charged species.  ANC is defined in chapter 2 (equation 2-11) as the difference
between strong anions and cations:

ANC = 2([CA2+] + [Mg2+]) + [K+] + [NH4+] - (2[SO4 2'] + [NCV] + [Cl'])               (7-1)

While the chemistry can be complex, the co-dependency between ANC  and pH is explained by
recognizing that positively charged hydrogen, H+, is incorporated in the charge balance
relationships related to the overall solution chemistry which also defines ANC.  The positive,
directional co-dependency (i.e.,  ANC and pH increase together) is further explained in concept
as ANC reflects how much strong acid (i.e., how much hydrogen ion) it takes to titrate to an
equivalence point.
       Strong observed correlations between pH and ANC (Fig 7-2) support these mechanistic
relationships.  There generally is an S-shaped relationship between pH and ANC, which suggests
that the linear part of the ANC and pH relationship is useful in guiding the discussion on ranges
of ANC (section 7.4) which relates most closely to adverse effects.   The asymptotic parts of the

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ANC-pH curve reflect the inherent buffering capacity of ANC as hypothetical additions of acid
would not result in significant pH change over those flat portions of the curve.
         300
         ICO*
           0-
         -100
         -100-
         -300
                    •   measured
                    *   calculated
Figure 1-2.
	       PH
 The relationship between pH and ANC under equilibrium conditions with mineral phase gibbsite.
 Triangles indicate calculated values while circles indicate measurements (Bi and Liu 2001).
•   What does the available evidence show concerning associations between ANC and
    aquatic acidification effects?
       As discussed in chapter 3, there are well established examples of ANC correlating
strongly with a variety of ecological effects which are summarized in Table 3-1.  Because pH
and ANC are well correlated (Figure 7-2) and linearly dependent over the pH ranges (4.5 - 6)
where adverse ecological effects are observed (Figure 3-3), evidence of clear associations exist
between ANC and adverse ecological effects (Figures 7-3 and 7-4). In large measure, these
figures, as well as the related effects discussions in chapter 3, speak directly to the
appropriateness of ANC with respect to its use as  an ecological indicator.
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20   0   20   40   60   BQ   K»  120  140   60
                                              14

                                              12-

                                              10-
                                           &
                                           .c
                                           V)
                                           il
                                           •s
                                                                I      I
                                                                100    200   300
                                                                ANC((ieq/L)
                                                                             I
                                                                            400
Figure 7-3 and 7-4. (Left) Relationship between ANC and number of fish species present in aquatic
freshwater ecosystems in Shenandoah National Park (Source:  Arthur Bulger, University of Virginia,
reproduced from NAPAP, 2005.).  (Right) Number offish species per lake versus acidity status,
expressed as ANC, for Adirondack lakes. The data are presented as mean (filled circles) and range (bars)
of species richness within 10 (ieq/L ANC categories, based on data collected by the Adirondack Lakes
Survey Corporation Source: Sullivan et al. (2006).
•  Why is ANC a more appropriate ecological indicator than pH?
       The previous two discussions established a clear association between ANC and
ecological effects, while acknowledging a more direct causal relationship between pH and
effects. Nonetheless, ANC is preferred as an ecological indicator based on its superior ability to
provide a linkage with deposition in a meaningful and quantifiable manner, a role that is served
far more effectively by ANC than by pH.  While both ANC and pH are clearly associated with
the effects of concern, ANC is superior in linking these effects to deposition.
       The basis for this conclusion is that acidifying atmospheric deposition of nitrogen and
sulfur is a direct input of potential acidity (ACY), or, in terms of ANC, such deposition is
relevant to the major anions in equation 7-1 that reduces the capacity of a water body to
neutralize acidity. Consequently, there is a well defined linear relationship between potential
acidifying deposition and ANC. This ANC-deposition relationship facilitates the linkage
between ecosystem models that calculate an ecological indicator and the atmospheric deposition
of NOy and SOx.  On the other hand, there is no direct linear relationship between deposition
and pH.  There certainly is a relationship, as acid inputs from deposition lower pH, but the
relationship can be extremely nonlinear and there is no direct connection from a modeling or
mass balance perspective between the amount of deposition entering a system and pH.  The term
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"mass balance" underlies the basic formulation of any physical modeling construct, atmospheric
or aquatic systems, and refers to the accounting of the flow of mass into a system, the
transformation to other forms, and the loss due to flow out of a system and other removal
processes. ANC is a conserved property. This means that ANC in a water body can be
accounted for by knowledge of how much ANC initially exists, how much flows in and is
deposited, and how much flows out.  In contrast, hydrogen ion concentration in the water, the
basis for pH, is not a conserved property as its concentration is affected by several factors such
as temperature, atmospheric pressure, mixing conditions of a water body, and the levels of other
several chemical species in the system. The disadvantage of pH lacking conservative properties
is that there is a very complex connection between changes in ambient air concentrations of NOy
and SOx and pH.
       The discussion of basic water chemistry of natural systems in chapter 2 provides further
details on why pH is not a conserved quantity and is subject to rapid transient response behavior
that makes it difficult to use as a reliable and functional ecological indicator. For now, we can
use the observed pH-to-ANC relationship (Figure 7-2) to partially explain the concern with pH
responding too abruptly. In the region where pH ranges roughly from 4.5 to 6 and is of greatest
relevance to effects (as seen in Figure 3-3),  there clearly is more sensitivity of pH to changes in
ANC in the ANC range from approximately 0 to 50 |ieq/L.  We focus on this part of the ANC-
to-pH relationship when we say that ANC associates well  with pH in a fairly linear manner.
However, the pH range from 4.5 to 6 also includes one of the very steepest parts of the slope
relating pH as a function of ANC (Figure 7-2), where ANC ranges down below 0 |ieq/L, which
is subject to very rapid change in ANC, or deposition inputs.  This part of the relationship
coincides with reduced levels of ANC and hence with reduced ability to neutralize acids and
moderate pH fluctuations.  This response behavior can be  extended to considering how pH
would change in response to deposition, or  ambient concentrations, of NOy and SOx, which can
be viewed as "ANC-like" inputs.
       In summary, because  ANC clearly links both to biological effects of aquatic acidification
as well as to acidifying inputs of NOy and SOx deposition, staff concludes that ANC is an
appropriate ecological indicator for relating adverse aquatic ecosystem effects to acidifying
atmospheric deposition of SOx and NOy, and is preferred to other potential indicators.  In
reaching this conclusion, we note that in its review of the first draft PA, CAS AC concluded that

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"information on levels of ANC protective to fish and other aquatic biota has been well developed
and presents probably the lowest level of uncertainty in the entire methodology" (Russell and
Samet, 2010a).  In its more recent review of the second draft PA, CASAC agreed "that acid
neutralizing capacity is an appropriate ecological measure for reflecting the effects of aquatic
acidification" (Russell and Samet, 2010b; p. 4).
7.2.2   Linking ANC to deposition.
         Linking atmospheric
         SOx and NOy deposition to
         ecological indicator
•  What does the available evidence show concerning the linkage of ANC to acidifying
   deposition?
       There is evidence to support a quantified relationship between deposition of nitrogen and
sulfur and ANC. This relationship was analyzed in the REA for two case study areas, the
Adirondack and Shenandoah Mountains, based on time-series modeling and observed trends.
Modeled long-term trends over time
       In the REA analysis, long-term trends in surface water nitrate, sulfate and ANC were
modeled using Model of Acidification of Groundwater in Catchment (MAGIC) for the two case
study areas. These data were used to compare recent surface water conditions (2006) with
preindustrial conditions (i.e. preacidification 1860). The results showed a marked increase in the
number of acid impacted lakes, characterized as a decrease in ANC levels, since the onset of
anthropogenic nitrogen and sulfur deposition (see chapter 2).
Observed recent trends over time
       In the REA, more recent trends in ANC, over the period from 1990 to 2006, were
assessed using monitoring data collected at the two case study areas. In both case study areas,
nitrate and  sulfate deposition decreased over this time period. In the Adirondack Mountains, this
corresponded to a decreased concentration of nitrate and sulfate in the surface waters and an
increase in  ANC (REA, section 4.2.4.2). In the Shenandoah Mountains, there was a slight
decrease in nitrate and sulfate concentration in  surface waters corresponding to modest increase
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in ANC from 50 ueq/L in 1990 to 67 ueq/L in 2006 (REA, section 4.2.4.3 and Appendix 4,
section 3.4).
•  What ecosystem modeling approaches should be considered to link ANC and
   deposition?
       In the REA, the quantified relationship between deposition and ANC was investigated
using ecosystem acidification models, also referred to as acid balance models or critical loads
models (chapter 2 above; REA, chapter 4 and Appendix 4).  These models quantify the
relationship between deposition of nitrogen and sulfur and the resulting ANC in surface waters
based on an ecosystem's inherent generation of ANC and ability to neutralize nitrogen
deposition through biological and physical processes.  A critical load is defined as the amount of
acidifying atmospheric deposition of nitrogen and sulfur beyond which a target ANC is not
reached. Relatively high critical load values imply that an ecosystem can accommodate greater
deposition levels than lower critical loads for a specific target ANC level.  Ecosystem models
that calculate critical loads form the basis for linking deposition to ANC.
       As discussed in chapter 2, both dynamic and steady state models calculate ANC as a
function of ecosystem attributes  and atmospheric nitrogen and sulfur deposition, and can be used
to calculate critical loads.  Steady state models are time invariant and reflect the long term
consequences associated with an ecosystem reaching equilibrium under a constant level of
atmospheric deposition.  Dynamic models are time variant and  take into account the time
dependencies inherent in ecosystem hydrology, soil and biological processes.  Dynamic models
like MAGIC can provide the time series response of ANC to deposition whereas steady state
models provide a single ANC relationship to any fixed deposition level.  Dynamic models
naturally are more complex than steady state models as they attempt to capture as much of the
fundamental biogeochemical processes as practicable, whereas  steady state models depend on far
greater parameterization and generalization of processes that is  afforded, somewhat, by not
having to accounting for temporal variability.
•  What is an appropriate modeling approach to link ANC  and deposition for
   development of a nationally applicable NAAQS?
       Steady state models are capable of addressing the question of what does it take to reach
and sustain a specific level of ANC, which is the  question most  relevant to the development of a
NAAQS. Dynamic  models are also capable of addressing that question, but can also address the

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question of how long it takes to achieve that result. In determining an appropriate modeling
approach for the development of a NAAQS, we consider both the relevance of the question
addressed as well  as the ability to perform modeling that provides relevant information for
geographic areas across the country.
       Dynamic models require a large amount of catchment level- specific data relative to
steady state  models. Because of the time invariant nature of steady state models, the data
requirements that  integrate across a broad spectrum of ecosystem processes is achievable and
available now at the national level.  In contrast, the data needs to support  dynamic models for
national-scale analyses simply are not available at this time.  Water quality data exist for
developing a national data base for modeling nearly 10,000 catchments in the contiguous U.S.  In
addition, the information provided by steady state modeling would be sufficient to develop and
analyze alternative NAAQS and the kind of protection they would afford.  While it would be an
important goal to  also obtain information about how much time it would take for a target ANC
level to be achieved, the absence of such information does not preclude developing and
evaluating alternative NAAQS using the AAI structure.  Based on the above considerations, staff
concludes that at this time steady state critical load modeling is an appropriate tool for linking
long-term ANC levels to atmospheric deposition of nitrogen and sulfur for development of an
AAI that has national applicability.
•  How does a steady state critical load model establish a linkage between ANC and
   associated levels of nitrogen and sulfur deposition?
       The steady state critical load model is used to define the amount of atmospheric
deposition of nitrogen (N) and sulfur (S) beyond which a target ANC is not achieved and
sustained. It is expressed as:
               = ([BC]o  - [ANClim])Q + Neco                               (7-2)
Where:
 CLANCiim(N + S) is the critical load of deposition, with units of equivalent charge/(area-time);
 [BC]o* is the natural contribution of base cations from weathering, soil processes and
       preindustrial deposition, with units of equivalent charge/volume;
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Q is the catchment level runoff rate governed by water mass balance and dominated by
       precipitation, with units of distance/time;
 [ANCurn] is the target ANC value, with units of equivalent charge/volume; and
Neco is the amount of nitrogen deposition that is effectively neutralized by a variety of biological
       (e.g., nutrient uptake) and physical processes, with units of equivalent charge/(area-time).

       Equation 7- 2 is a modified expression that adopts the basic formulation of the SSWC
and FAB steady state models that are described in chapter 2. More detailed discussion of the
rationale, assumptions and derivation of equation 7- 2, as well as all of the equations in this
chapter, are included in Appendix B.  For now, the equation simply reflects the amount of
deposition, CLANCiim(N + S), associated  with a sustainable long-term ANC target, [ANCum], given
the natural system ANC generation, [BC]o, and the capacity of the natural system to neutralize
nitrogen deposition, Neco. We note that this expression of critical load is valid when nitrogen
deposition is greater than Neco. The runoff rate, Q, allows for balancing mass in the two
environmental mediums - atmosphere and catchment.
       In considering the contributions  of SOx or NOy species to acidification, it is useful to
think of every depositing nitrogen atom as supplying one equivalent charge unit and every sulfur
atom as depositing two charge units.  We use equivalent charge per volume as a normalizing tool
in place of the more familiar metrics such as mass or moles per volume.  This allows for a clearer
explanation of many of the relationships between atmospheric and ecosystem processes  that
incorporate mass and volume unit conventions somewhat specific to the environmental media of
concern (e.g., m3 for air and liter for liquid water).  Equivalent charge reflects the chemistry
equilibrium fundamentals that assume electroneutrality, or balancing charge where the sum of
cations always equals the sum of anions.  This fundamental relationship is behind, but only
partially explains, the simple ANC equation (7-1) introduced earlier:

       ANC = sum of major cations - sum of major anions.

       At this stage we use the terms S  and N in the CLANCum (N + S) term to broadly represent
all species of sulfur or nitrogen that can contribute acidifying deposition. This follows
conventions used in the scientific literature that addresses critical loads, and it reflects all

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possible acidifying contributions from any sulfur or nitrogen species. For all practical purposes,
S reflects SOx as described in section 7.1, the sum of sulfur dioxide gas and particulate sulfate.
However, N includes both oxidized forms, consistent with the ambient  indicator, NOy, in
addition to reduced nitrogen species, ammonia and ammonium ion, referred to as NHx.  NHx is
included in the critical load formulation because it contributes to potentially acidifying nitrogen
deposition.  Consequently, from a mass balance or modeling perspective, the form of the
standard must account for NHx as described below. The data requirements for equation 7-2 are
addressed later in section 7.2.5 after we complete the discussion of form.
•  How is reduced nitrogen deposition, NHx, considered separately from oxidized forms of
   nitrogen?
       Equation 7-2 relates total nitrogen deposition to ANC. However,  for the AAI form of the
standard it is important to separately identify and include the direct relationship between ambient
air indicators of oxides of nitrogen and sulfur and ANC, as illustrated in Figure 7-1.  This can be
done by separating total nitrogen deposition, N, into oxidized, NOy, and reduced, NHx,
components:
               NHx + SOx) = ([BC]o - [ANCUm])Q + Neco                         (7- 3)
We can define a critical load in terms of NOy and SOx that takes into account the available
supply of NHx deposition:

CLANCtim(NOy + SOx) = ([BC]o  - [ANCtim])Q + Neco - NHx                          (7-4)

Where NHx represents the combined wet and dry deposition of ammonia, NET?, and ammonium
ion, NH4.  By separating out NHx deposition from the aggregated critical load, the amounts of
combined NOy and SOx deposition for the critical load are identified.
7.2.3   Linking deposition to allowable concentrations
         Linking deposition to "allowable"
         concentrations of ambient air
         indicators of oxides of N and S
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       The last major component of the simplified form illustrated in Figure 7-1 addresses the
linkage between deposition and allowable concentrations of ambient air indicators, NOy and
SOx.
•  How do we link deposition to allowable concentrations?
       To link ambient air concentrations with deposition, we define a transference ratio, T, as
the ratio of total wet and dry deposition to concentration, consistent with the area and time period
over which the standard is defined.   Since we intend to express deposition of NOy and SOx in
terms of NOy and SOx concentrations, we define two transference ratios:

Tsox =  Dep(SOx)/[SOx], and
TNOy = Dep(NOy)/[NOy]

Where;
 Dep(SOx or NOy) is the combined dry and wet deposition of SOx or NOy, and
[SOx or NOy] is the ambient air concentration of SOx or NOy.

Before discussing the rationale, assumptions and information to develop the transference ratios,
we reconstruct equation 7-4 in concentration terms using transference ratios:

CLANcumfNOy + SOx) = ([BC]o - [ANCUm])Q + Neco - NHx                          (7-4)

Consider the CLANCUm(NOy  +  SOx) term as representing the combinations of NOy and SOx
deposition that would meet  a critical load:

Dep (NOy) + Dep(SOx) = ([BC]o - [ANCUm])Q + Neco - NHx                       (7-5)

Express the deposition of NOy and SOx in concentration terms and  rewrite (7-5):

[NOy]TNOy + [SOx]TSOx = ([BC]o  - [ANCUm])Q + Neco - NHx                        (7-6)

Rearrange to define critical  load in terms of NOy and SOx concentration;
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CL(N+S) = ([BC]o  - [ANCumDQ + Neco = [NOy]TNOy + [SOx]TSOx + NHx           (7-7)

Equation 7-7 traces back to the original critical load expression, equation 7-2, with two
refinements: depositions are translated to ambient air concentrations through the transference
ratios, and reduced nitrogen deposition, NHx, is separated from total nitrogen deposition to allow
for an expression that relates the ambient air indicators, NOy and SOx, and ANC. The rationale
underlying transference ratios follows.
•  What approaches are considered for developing transference ratios?
       Transference rations are a modeled construct, and therefore we are not able to compare
directly these ratios with explicit measurable quantities.  There is an analogy to deposition
velocity,  as a transference ratio is basically an aggregated weighted average of the deposition
velocities of all contributing species across dry and wet deposition, and transference ratio units
are expressed as distance/time. However, wet deposition commonly is not interpreted as the
product of a concentration times a velocity.  Direct wet deposition observations are available
which integrate all of processes, regardless of how well they may be understood, related to wet
deposition into a measurable quantity.  There has been a history of nomenclature and
conventions using terms such as washout ratios that incorporate the essence of transferring
ambient mass to rain and cloud droplets, as summarized in Seinfeld and Pandis (1998).  There
are reasonable analogies between the processes governing dry and wet deposition, from a
fundamental mass transfer perspective.  In both cases there is a transfer of mass between the dry
ambient phase and another medium, either a surface or vegetation in the  case of dry deposition,
or a rain droplet or cloud in the case of wet precipitation. The specific  thermodynamic properties
and chemical/biological reactions that govern the transfer of dry mass  to plants or aqueous
droplets differ, but either process can be based on conceptualizing the  product of a concentration,
or concentration difference, times a mass transfer coefficient which is analogous to the basic dry
deposition model: dry deposition = concentration x velocity. Indeed, Seinfeld and Pandis (1998)
utilize the concept of a wet  deposition velocity in explaining wet deposition processes, and this
rationale is captured in more detail in Appendix F.
       Transference ratios require estimates of wet deposition (NOy and SOx), dry deposition
(NOy and SOx) and concentrations of NOy and SOx.  Possible sources of information include
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model estimates or a combination of model estimates and observations, recognizing that dry
deposition is a modeled quantity that can use observed or modeled estimates of concentration.
The limited amount of NOy measurements in acid sensitive areas as well as the combination of
representative NOy, 862 and 864 observations generally preclude the use of observations for a
standard that is applicable nationally.
       One could consider  a blending of observations and models to take advantage of their
relative strengths, for example, combining the NADP wet deposition observations, modeled dry
deposition, and a mix of modeled  and observed concentrations, using the model for those species
not measured or measured with very sparse spatial coverage.   A potential disadvantage of
mixing and matching model estimates is to lose consistency afforded by using just modeling
alone.  A modeling platform like CMAQ is based on adhering to consistent treatment of mass
conservation, by linking emission inputs with air concentrations and concentrations to
deposition. Inconsistencies from combining processes from different analytical platforms
increase the  chance that mass (of nitrogen or sulfur) would unintentionally be increased or
decreased as the internal checking that assures  mass conservation is lost.  Transference ratios
incorporate a broad suite of atmospheric processes and consequently  an analytical approach that
instills consistency in the linkage of these processes is preferable to an approach lacking such
inherent consistency. This contention does not mean that observations alone, if available, could
not be used,  but suggests that the inconsistencies in combining models and observations for the
purposes of developing transference ratios has the potential for creating unintended artifacts.
       While there is a reasonable conceptual basis for the concept of an aggregated deposition
velocity we are calling a transference ratio, there is very limited ability to compare observed and
calculated ratios.  This is because  the deposition velocity is dependent on individual species, and
the mass transfer processes  of wet and dry removal, while conceptually similar, are different.
Consequently, there does not exist a meaningful approach to measure such an aggregated or
lumped parameter. Therefore,  at this time our evaluation of transference ratios is based on
sensitivity studies, analysis  of variability, and comparisons with other models, as described in
Appendix F.
       The interannual variability, as well as the sensitivity to emission changes of roughly 50%,
result in changes of transference ratios of approximately  5 - 10%. Part of the reason for this
inherent stability is due to the co-dependence of concentration and deposition. For example, as

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concentrations are reduced as a result of emissions reductions, deposition in turn is reduced since
deposition is a direct linear function of concentration leading to negligible impact on the
deposition-to-concentration ratio. The same line of reasoning explains why an overestimate of
concentration likely does not induce a bias in the transference ratio.  While it is important to
continue to improve the model's ability to match ambient concentrations in time and space, the
bias of a modeled estimate of concentration relative to observations does not necessarily result in
a bias in a calculated transference ratio.  In effect, this consideration of bias cancellation reduces
the sensitivity of transference ratios to model uncertainties and affords increased confidence in
the stability of these ratios. Based on the series of sensitivity and variability analyses, staff
concludes that the transference ratios are relatively stable and provide a sound metric for linking
deposition and concentration in the form of the standard.
       Transference ratios are dependent on the platform they are constructed upon.
Comparisons of transference ratios constructed from different modeling platforms do exhibit
significant differences.  While this divergence of results may be explained by a variety of
differences in process treatments, input fields and incommensurabilities in species definitions
and spatial configurations, it does suggest two very important conclusions. First, the idea of
using multiple platforms for different parts of the country may be problematic as there does not
exist a reliable approach to judge acceptance which is almost always based on comparisons to
observations. Second, since transference ratios  are based on concentrations and deposition, as
the uncertainties in each of those components are reduced, the relative uncertainty in the ratios
also is reduced.  This means that basic improvements in the model's ability to reproduce
observed wet deposition and ambient concentration fields enhance the relative  confidence in the
constructed transference ratios. Similarly, as in-situ dry deposition flux measurements become
available that enable a more rigorous evaluation and diagnosis of modeled dry  deposition
processes, the expected improved treatment of dry deposition also would increase confidence in
transference ratios.  Finally, deposition is directly related to ambient air concentrations.  Models
like CMAQ rely on the concentration-to-deposition linkage to calculate deposition, which is the
foundation for broadly based and robust assessments addressing atmospheric deposition. In
principle, the use of a modeled constructed transference ratio is based on the same premise by
which we use models to estimate deposition in the first place.
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       The shortage of widely available ambient air observations and the fact that estimates of
dry deposition requires modeling, collectively suggests that a unified modeling platform is the
best approach for constructing transference ratios.  Staff has considered CMAQ and other models
(see chapter 2), such as CAMx and the Canada's AURAMS - A Unified Regional Air-quality
Modeling System (Smythe et al, 2008), and concludes that CMAQ is the preferred modeling
platform for constructing transference ratios for purposes of this NAAQS review.  This
conclusion reflects our view that for the purposes of defining transference ratios,  a modeling
platform should (1) be a multiple pollutant model recognizing the myriad of connections across
pollutant categories that directly and indirectly impact nitrogen and sulfur characterization, (2)
include the most comprehensive scientific treatments of atmospheric processes that relate
directly and indirectly to characterizing concentrations and deposition, (3) have an infrastructure
capability that accommodates the inclusion of improved scientific treatments of relevant
processes and important input fields, and (4) undergo frequent reviews regarding the adequacy of
the underlying science as well as the appropriateness in applications. The CMAQ platform
exhibits all these characteristics. It has been (and continues to be) extensively evaluated for
several pollutant categories,  is supported by a central infrastructure of EPA scientists, with
considerable interfacing with the scientific research communities in academia and industry,
whose mission is to improve and evaluate the CMAQ platform.  More directly, CMAQ, and its
predecessor versions, has a long track record going back to the NAPAP in the 1980's of specific
improvements in deposition  processes, which are described in Appendix F.
7.2.4   Completing the link from ecological indicator to ambient air indicators
                        Linking atmospheric
                        oxides of S and N deposition to
                        ecological indicator
  Ecological effects and                                  Linking deposition to "allowable"
  ecological indicator                                    concentrations of NOy and SOx
  (ANC)
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•  How is a target long-term ANC level linked to appropriate terms of ecosystem
   attributes, reduced nitrogen deposition, and ambient air indicators?
       The two previous sections described the links between long-term ANC and deposition
(7.2.2) and deposition and ambient air concentration (7.2.3) provided by equation 7-7:

CL = ([BC]o  - [ANClim])Q + Neco = [NOy]TNOy + [SOx]TSOx + NHx                  (7-7)

Equation (7-7) represents a relationship that defines the ambient air concentrations of NOy and
SOx that would not exceed a specified  critical load.  The terms on the right side of equation 7-7
represent the maximum amount of acidifying deposition, expressed in terms of the
concentrations of ambient air indicators, NOy and SOx, as well as the deposition of NHx, that
would not exceed a specified critical load.  The difference between actual total acidifying
deposition and the critical load is referred to as "exceedance deposition," DEPex, where:

DEPex = [NOy]TNOy + [SOxJTSOx + NHx - CL                                      (7-8)

Thus, exceedance deposition is the amount of acidifying deposition in excess  of the amount of
deposition that would just achieve a specified critical load.

We define a related term,"ANClim  exceedance," ANClimex, that is directly proportional to
deposition exceedance by dividing by the runoff rate, Qr, representative of the area over which
all the atmospheric terms are defined:

ANClimex = DEPex/Qr = {[NOy]TNOy + [SOx]TSOx + NHx - CL}/Qr                   (7-9)

ANClim exceedance (ANClimex) is the difference between a target ANC (ANClim) and a
calculated ANC for an area. Thus, we  can calculate an ANC value (ANCcalc) by the following
equation:

ANCcalc = ANClim - {[NOy]TNOy + [SOxJTSOx + NHx - CL}/Qr                      (7-10)
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The terms in equation 7-10 are then rearranged to highlight the connection from the ecological
indicator, ANC, to ambient air indicators, NOy and SOx, in terms of a representative critical load
(CLr) for an area:

ANCcalc = {ANClim + CLr/Qr} - NHx/Qr - TNOy [NOy]/Qr - TSOx[SOx]/Qr             (7-11)

       Equation 7-11 is the basic expression of the standard which translates the simple
conceptual diagram into an explicit expression that relates ANC as a function of the ambient air
indicators, NOy and SOx.  Based on equation 7-11, we define an aquatic acidification index
(AAI) that is more simply stated in terms that emphasize the ambient air indicators:

       AAI = F1-F2- F3[NOy] - F4[SOx]                                      (7-12)

where the AAI represents the long term (or steady state) ANC level associated with ambient air
concentrations of NOy and  SOx.   The AAI is the potential the atmosphere affords in influencing
ecosystem ANC.  The factors Fl through F4 convey three attributes:  a relative measure of the
ecosystem's ability to neutralize acids (Fl), the acidifying potential of reduced nitrogen
deposition (F2), and the deposition-to-concentration translators for NOy (F3) and SOx (F4).
Specifically:
Fl = ANClim + CLr/Qr;
F2=  NHx/gr = NHx deposition divided by Qr;
F3 =  TNoy/ Qr; TNoy is the transference ratio that converts deposition of NOy to ambient air
       concentrations of NOy; and
F4 =  TSOX/ Qr; TSOX is the transference ratio that converts deposition of SOx to ambient  air
       concentrations of SOx.
All of these factors include  representative Qr to maintain unit (and mass) consistency between
AAI and the terms on the right side of equation 7-12.

       We note that the Fl  factor incorporates an ecosystem's ability to generate acid
neutralizing capacity through base cation supply ([BC]*o) and to neutralize acidifying nitrogen
deposition through Neco, both of which are incorporated in the CL term. Because Neco can only
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neutralize nitrogen deposition (oxidized or reduced) there may be rare cases where Neco exceeds
the combination of reduced and oxidized nitrogen deposition.  Consequently, to ensure that the
AAI equation is applicable in all cases that may occur, we recognize that equation 7-12 is
conditional on total nitrogen deposition, (NHx + F3[NOy]}, being greater than Neco.  In rare
cases where Neco is greater than {NHx + F3[NOy]}, F2 are F3 would be set equal to 0.  In such
cases, CLr would be defined only in terms of acidifying deposition of sulfur:

CLr(S) = ([BC]o - [ANCumDQ, only when Neco^L {NHx + F3[NOy]}                   (7-13)

The consequence of setting F2 and F3 to zero and eliminating Neco from the CL expression is
simply to constrain the AAI calculation just to SOx as nitrogen would have no bearing on
acidifying contributions in this case.
       Staff concludes that equation 7-12, which defines an AAI, is ecologically relevant and
appropriate for use as the form of a national standard designed to provide protection for aquatic
ecosystems from the  effects association with acidifying deposition associated with
concentrations of oxides of nitrogen and sulfur in the ambient air. We note, however, that
equation 7-12 does not, in itself, define the spatial areas over  which the terms of the equation
would apply.  To specify values for factors Fl through F4, it is necessary to define spatial areas
over which these factors are determined.  Thus, it is necessary to identify an approach for
spatially aggregating water bodies into ecologically meaningful regions across the U.S., as
addressed in the next section.
7.2.5  Spatial Aggregation
       One of the unique aspects of this review is the need to consider the spatial areas over
which values for the factors in the AAI equation that defines the form of the standard are
quantified. Ecosystems across the U.S. exhibit a wide range of geological, hydrological and
vegetation characteristics that influence greatly the  ecosystem parameters, Q, BCo* and Neco that
are incorporated in the AAI. Variations in ecosystem attributes naturally lead to wide variability
in the sensitivities of water bodies in the U.S. to acidification, as well as in the responsiveness of
water bodies to changes in acidifying deposition.  Consequently, variations in ecosystem
sensitivity must be taken into account in developing a national standard.  In developing a
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national secondary standard to protect public welfare, our focus is on protecting sensitive
populations of water bodies, not on each individual water body, which is consistent with our
approach to protecting public health through primary standards that focus on susceptible
populations, not on each individual.
       In this section, we first describe alternative approaches to  defining ecologically relevant
regions across the U.S.  Once spatially aggregated regions are established, we then consider
approaches to characterizing each region as acid sensitive or relatively non-acid sensitive based
on alkalinity and ANC data. This characterization facilitates a more detailed analysis of those
regions that are relatively  more acid sensitive.  We also use this characterization to avoid over-
protection in relatively non-acid sensitive regions that would receive limited benefit from
reductions  in the deposition of oxides of nitrogen and sulfur with  respect to aquatic acidification
effects. Further, we discuss approaches to developing representative values  of each of the terms
in the AAI equation (factors Fl through F4) for each ecologically relevant region.  These
following sections generally address spatial aggregation approaches applicable to the United
States. The approaches discussed below, however, are limited to  the contiguous United States
since there is insufficient data available for Hawaii, Alaska and the territories to consider
applying such approaches at this time. Other approaches to apply to these areas are discussed
below.
       Stated more simply, this section discusses appropriate ways to divide the country into
ecologically relevant regions; to characterize each region as either acid sensitive or relatively
non-acid sensitive; and to determine values of factors Fl  through  F4 for each region, taking into
consideration the acid sensitivity of each region. For each such region, the AAI would be
calculated based on the values of factors Fl through F4 specified  for that region.
Approaches to spatial aggregation
       In considering approaches to spatial aggregation,  staff focused on methods that have been
developed to define ecologically relevant regions, referred to as ecoregions,  which are
meaningfully related to the factors that are relevant to aquatic acidification.  As noted above,  we
did not focus on looking at each individual water body. We first considered the broadest
aggregation possible that looked at the entire nation as one region. We recognize that
aggregating over the entire nation would preclude taking into account the inherent variability in
atmospheric and ecological factors that fundamentally modify the relationships that are central to
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the development of an ecologically relevant AAI.  As a consequence, we conclude it is
appropriate to consider approaches that divide the country into ecologically relevant regions for
the purpose of defining appropriate spatial areas over which AAI factors would be specified and
the AAI would be calculated.
•   What approaches are available to define ecologically relevant regions in the U.S.?
       Ecoregions are areas of similarity regarding patterns in vegetation, aquatic, and terrestrial
ecosystem components.  Available ecoregion categorization schemes include EPA's Omernik
classifications (http://www.epa.gov/wed/pages/ecoregions.htm. Omernik, 1987), the National
Ecological Observatory Network (NEON, http://www.neoninc.org/) domains, and Baily's
ecoregions developed for the United  States Forest Service (USFS).
       The NEON domains are under development and the current design is based on 20 eco-
climatic regions, each with similarities in vegetation,  landforms, and climate.  One goal of
NEON, which is supported by the  U.S. National Science Foundation, is to develop a baseline of
ecological data for use in variety of applications, especially to observe the effects of changing
climates on ecosystem parameters  and performance.  Bailey's ecoregions also group regions
based on similar vegetation and climatic conditions.  There are no apparent  advantages of NEON
or Bailey's scheme relative to the Omernik classification system. The lack of more resolved
spatial groupings as well as being in the developmental stage limits the utility of the NEON
domains at this time.  Neither Bailey's scheme nor NEON has undergone a level of peer review
and scientific scrutiny comparable to that achieved with the Omernik scheme. Omernik's
scheme is well documented and used frequently in the ecosystem community which has resulted
in readily accessible data and an increased knowledge base of its utility.
       Omernik's ecoregions are categorized using a holistic, "weight-of-evidence" approach in
which the relative importance of factors may vary from region to region. The method used to
map ecoregions is described in Omernik (1987), as one that is:

       ". . . based on the premise that ecological regions can be identified through the
       analysis of the patterns and the composition of biotic and abiotic phenomena that
       affect or reflect differences in ecosystem quality and integrity (Omernik, 1987).
       These phenomena include geology, physiography, vegetation, climate, soils, land
       use, wildlife, and hydrology. The relative importance of each characteristic varies
       from one ecological region to another regardless of the hierarchical level."
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       The first publication of the ecoregions based on Omernik's weight-of-evidence approach
was published in 1987 (Omernik, 1987).  Current maps found in
http ://www. epa. gov/wed/pages/ecoregions. htm are refinements and revisions of the 1987
publication. Hierarchical levels were developed and a Roman numeral classification scheme was
adopted to distinguish coarser (more general) and finer (more detailed) categorization. Level I is
the coarsest level, dividing North America into 15 ecoregions . At level II, the continent is
subdivided into 52 ecoregions . Level III is a further subdivision of level II, and divides North
America into 120+ ecoregions. Level IV is a subdivision of level III, and development of maps
for this level is currently in progress.
•   What ecoregion categorization scheme is most applicable for the purpose of defining the
    AAI?
       For the reasons discussed above, staff concludes that Omernik's ecoregion classification
is the most appropriate method to consider for  the purposes of this review as it offers several
levels of spatial delineation, has undergone an  extensive scientific peer review process, and has
explicitly been applied to delineating acid sensitive areas within the United States. Further, we
conclude that ecoregion level III (Figure 7-5) resolution with 84 defined regions in the
contiguous United States5 is the most appropriate level to consider for this purpose. The spatial
resolution afforded by Level III strikes an appropriate balance relative to the reasoning that
supports staff conclusions on indicators, as discussed above in section 7.1. We conclude that the
most detailed level of resolution (level IV) is not appropriate given (1) the limited data
availability to address nearly  1000 subdivisons within that level and (2) the currently evolving
nature of level IV regions.  Further,  we conclude that level III regions are preferred to level II in
that level III regions, but not level II regions, are largely contiguous in space which allows for a
more coherent development of information to quantify the AAI factors and to  characterize the
concentrations of NOy and SOx in the ambient air within each region.
       Appendix C includes a description of each level III ecoregion. While the use of
ecoregions is an appropriate spatial  aggregation scheme for this NOy/SOx standard focused  on
aquatic acidification, many of the same ecoregion attributes may be applicable in subsequent
NAAQS reviews that may address other deposition-related aquatic and terrestrial ecological
effects.  Because atmospheric deposition is modified by ecosystem attributes, the types of
5 We note that an 85th area within Omernik's Ecoregion Level III is currently being developed for California.
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vegetation, soils, bedrock geology, and topographic features that are the basis of this ecoregion
classification approach also will be key attributes for other deposition-related effects (e.g.,
terrestrial acidification, nutrient enrichment) that link atmospheric concentrations to an aquatic or
terrestrial ecological indicator. Just as this aquatic acidification standard links atmospheric and
ecosystem processes, future NOy/SOx standards that consider other deposition-related effects
may well include ecosystem-based processing of deposition inputs as they translate to a defined
ecological indicator.
                                           Omernik Ecoregion II Index Map
                                        I   | MISaSSIPPIALUVI^ANDSOLTHEASTUSACOASTAL PLAINS I   I TEXAS-LOLISIANA COASTAL PLAIN
                                        |   | MIXED WOCO PLAINS                        |   | UPPER GILA MOUNTAINS
                                        |   | MIXED 'WOCO SHIELD                        |   | '.WRM DESERTS
                                        |   | aZARK>DLAOUTA-AP=*LACI- IAN FORESTS            |   | WEST-CENTRAL SEMI-ARID PRAIRIES
                                        |   | 30JT-I CENTRAL SEMI-IRIDPRAIRE3              |   | WESTERN CCRDLLERA
                                        I   | SO JTHEASTERN USA PLAINS                   I   I WESTERN SIERRAMADREP1EDMONT
                   |   | MAS INS WEST COAST =OTEST |   | TAMAJLPAS-TEXASSEHIASIDPLAK
                   |   | MEDITERRANEAN CALFORNIA |   | TEMPERATE PRAIRIES
Eco_Level_lll_US
NA_L2MAME
|   | ATLANTIC HIGHLANDS
|   | CENTRALUSA,PLAINS
|   | CCLD DESERTS
I   | EVERGLADES
Figure 7-5.  Omernik Ecoregion II areas with ecoregion III subdivisions (http://www.epa.gov/wed/pages/ecoregions).
There are 20 Ecoregion II categories, each of which are further subdivided into a total of 84 Level III categories.
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Characterizing ecoregion level III sensitivity
       Staff has used Omernik' s original alkalinity data (chapter 2) and more recent ANC data
(as discussed above in chapter 2) to delineate two broad groupings of ecoregions:  acid sensitive
and relatively non-acid sensitive ecoregions. This delineation was performed to enable greater
focus on those regions with water bodies that generally have greater acid sensitivity and to avoid
over-protection in regions with generally less sensitive water bodies.  Our approach to
delineating acid sensitive and relatively non-acid sensitive regions, which is discussed more fully
below in section 7.5, included an initial numerical-based sorting scheme using ANC data.
Following this initial delineation, we reviewed other water quality parameters to identify
naturally acidic conditions associated with low base cation supply or high organic acid levels,
which would support characterizing a region as relatively less-acid sensitive, which is addressed
later in section 7.5.  In addition, we considered the degree to which ecoregions exceed
representative critical loads based on current and future deposition levels, which provided insight
into the likelihood that a region is naturally acidic and unlikely to be responsive to changes in
concentrations of NOy and SOx in the ambient air and thus to changes in related deposition
levels. These reviews based on analyzing available data were supplemented by considering to
what extent a region is characterized as a relatively pristine, rural undeveloped area that is not
predominantly managed for agricultural or forest products, as described in chapter 1.  This last
consideration allows for the application of common sense judgments to avoid over-protection,
which cannot be arrived at through data analysis alone. The overall objective is to produce a
logical and practical grouping of ecoregions that experience adverse conditions with respect to
aquatic acidification and are likely to respond to changes in concentrations of NOy and SOx in
the ambient air and to the related deposition levels.
        The initial delineation of acid sensitive and relatively non-acid  sensitive regions used
ANC data to determine the number of water bodies within the region with long-term ANC values
suggestive of acid sensitivity, so as to screen out regions with an overabundance of high ANC
values.  In our review of this ANC data, we identified regions that have greater than 5% of water
bodies with data with ANC values less  than 200 jieq/L and that have greater than 1% of water
bodies with ANC values less than 100 jieq/L. Applying these criteria yielded 29 acid sensitive
ecoregions (Figure 7-6). The resulting  acid sensitive ecoregions resemble the patterns of acid
sensitivity in the original Omernik alkalinity map (Figure 2-39) and in the similar ANC map
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(Figure 2-40), which is an expected outcome as Figure 7-6 is derived from ANC and ALK data.
The patterns reveal the simple observation that collection of ANC data has been targeted to areas
of suspected or known acid sensitivity.
       In addition, the acid sensitive ecoregions generally are characterized as areas with
mountainous, high elevation terrain or water bodies in Northern latitudes (Northern areas of
Minnesota, Wisconsin and Michigan; and New England).  The northern non-mountainous
regions share attributes (growing season, vegetation, soils and geology) similar to mountainous
regions and typically are located in rural areas, often in tracts of designated wilderness, park and
recreation areas.  Of the 29 acid sensitive ecoregions, the following six ecoregions are located in
two Level II ecoregions (i.e., Southeastern Plains (8.3) and Mississippi Alluvial and Southeast
Coastal Plains (8.5); Figure 7-6) and are characterized by relative lowland plains or transitional
lands between plains and hills:  Piedmont (8.3.4), Southeastern Plains (8.3.5), South Central
Plains (8.3.7), Middle Atlantic Coastal Plain (8.5.1), Southern Coastal Plains (8.5.3) and Atlantic
Coastal Pine Barrens (8.5.4). These coastal plains and transition areas are noted here and are
discussed in more detail below in section 7.5 in considering alternative criteria for delineating
acid sensitive and relatively non-acid sensitive ecoregions.
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     Alkalinity
        ] Less than 100 |jeq/L
     EcoRegions III
          | Non - Sensitive
          | Sensitive
           Likely Sensitive
     Surface Water
       =   Less than 100 ANC |jeq/L
                               Ecoregion Acid-Sensitivity
      Legend
        	I Sensitive Ecoregions
        _J Relatively Non-Sensitive Ecoregions

Figure 7-6.  The top panel captures some of the data details used in delineating acid sensitive and relatively non-
acid sensitive regions, which are shown in the bottom panel.  The red areas in the top panel reflect acid sensitive
areas with small samples sizes (less than 20) of water quality data.  The four cross-hatched acid sensitive regions in
the bottom panel are low elevation coastal plains type areas as discussed in sections 7.2 and 7.5.
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Establishing representative factors for the AAI equation
       Having concluded that the Omernik level III ecoregions are an appropriate approach to
spatial aggregation for the purpose of an aquatic acidification standard, we use those ecoregions
to define each of the factors in the AAI equation (equation 7-12 developed above in section
7.2.4):

AAI =  F1-F2- F3[NOy] - F4[SOx]                                                 (7-12)

Where:
Fl =  ANClim + CLr/Qr;
F2 =  NHx/Qr; NHx is the deposition divided by Qr;
F3 =  TNoy/Qr; TNoy is the transference ratio that converts deposition of NOy to ambient air
       concentrations of NOy; and
F4 =  TSox/Qr; TSOX is the transference ratio that converts deposition of SOx to ambient air
       concentrations of SOx.

The factors Fl through F4 in equation 7-12 are defined for each ecoregion by specifying
ecoregion-specific values for  each factor based on monitored or modeled data that are
representative of each ecoregion.  The Fl factor is also defined by a target ANC value
(ANClim), as discussed below in section 7.4.
       To specify ecoregion-specific representative values for the term CLr in factor Fl, we first
create a distribution6 of calculated critical loads for the water bodies in the ecoregion  for which
we have sufficient water quality and hydrology data.7 We then define the representative critical
load to be a specific percentile, the nth percentile, of the distribution of critical loads in the
ecoregion. Thus, for example, using the 90th percentile means that within an ecoregion, 90
percent of the water bodies would be expected to have higher calculated critical  loads than the
representative critical load. The choice of an appropriate range of percentile values to consider is
discussed below in response to the next question.  To specify ecoregion-specific representative
6 In this PA, the distribution of critical loads was based on CL values calculated with Neco at the lake level.
Consideration should be given to using a distribution of CLs without Neco and adding the ecoregion average Neco
value to the nth percentile critical load. This would avoid cases where the lake level Neco potentially could be
greater than total nitrogen deposition.
7 We judge the data to be sufficient for this purpose if data are available from more than 10 water bodies in an
ecoregion.
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values for the term Qr, which is used in factors Fl through F4, we use the median value of the
distribution of Q values that are available for water bodies within each ecoregion.
       To specify ecoregion-specific representative values for the remaining terms in the AAI
equation, NHx, TNoy, TSox, NOy, and SOx, we use data averaged over the ecoregion. Each of
these terms is based on 2005 CMAQ model simulations over  12-km grids, discussed above in
chapter 2. The CMAQ simulation provides estimates of deposition of NHx, NOy, and SOx, as
well as ambient air concentrations of NOy and SOx.  All of these terms are based on annual
average model outputs for each grid cell, which are spatially averaged across all the grid cells
contained in each ecoregion to calculate a representative annual average value for each
ecoregion.  The transference ratios, TNOY and TSOX, are calculated as the annual deposition
spatially averaged across the ecoregion and divided by the annual ambient air concentration
spatially averaged across the ecoregion.  We conclude that this approach of using spatially
averaged values is appropriate, largely due  to the relatively rapid mixing of air masses due to
gaseous-based advection and dispersion processes that typically results  in relatively
homogeneous air quality patterns for regionally dispersed pollutants.  In addition, there is greater
confidence in using spatially averaged modeled atmospheric fields than in using modeled point-
specific fields.
       Of these terms, NHx deposition perhaps exhibits greater spatial variability, as well as
overall uncertainty, than the other terms. On this basis, we conclude that it would also be
appropriate to consider allowing the use of alternative approaches to specifying the value of
NHx. One such approach might involve the use of more localized and/or contemporaneous
modeling in areas where this term is likely to be particularly variable and important.  Such an
alternative approach could allow for more localized changes over time in the concentration of
NHx to be reflected in the calculated AAI value for an ecoregion.  Other approaches might
involve the use of monitored NHx data as the concentration variable applied in dry deposition
modeling.
       The ecoregion-specific values for factors Fl through F4 would be codified as part of a
standard that is defined in terms of the AAI. For the purpose  of calculating AAI values that
reflect recent air quality  in the absence of NOy and SOx monitoring data in each ecoregion, we
use the annual average NOy and SOx concentrations that result from the 2005 CMAQ model
simulation. For the purpose of applying a standard defined in terms of the AAI in the future,

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NOy and SOx would be determined by measuring the annual average concentrations of NOy and
SOx in the ambient air.  The measured values of NOy and SOx would then be used in equation
7-12, together with the values for Fl through F4 that are specified for each ecoregion, to
calculate an annual AAI for an ecoregion.
•   What range of nth percentile values is appropriate to consider as part of the definition
    of the form of the standard?
       The nth percentile value chosen as part of the definition of the form of the standard is an
important parameter that directly impacts the representative critical load specified for each
ecoregion, and therefore the degree of protectiveness of the standard. A higher percentile
corresponds to a lower critical load and, therefore, to lower allowable ambient air concentrations
of NOy and SOx and the related deposition to achieve a target AAI level.  In conjunction with
specifying the values of factors Fl  through F4 as discussed  above, alternative forms for
consideration can be appropriately characterized in part by identifying a range of alternative
percentile values for consideration. Consequently, we assess alternative standards below
(section 7.5) by specifying alternative combinations of percentile values, as discussed here, and
target ANC values, which would equate to standard levels, discussed below in section 7.4, within
the ranges identified as appropriate to consider in this review.
       In identifying percentile values that are appropriate to consider, we take into
consideration the characterization of the ecoregions as acid  sensitive or as relatively non-acid
sensitive,  as discussed above. In considering ecoregions characterized as acid sensitive, we
judge that it is appropriate to  focus on the upper part of the distribution of critical loads, so as to
ensure that the ecoregion would be represented by relatively more acid sensitive water bodies
within the ecoregion.  Specifying the form in this way would help to define a standard that would
be protective of the population of acid sensitive water bodies within an ecoregion, recognizing
that even ecoregions characterized as acid  sensitive may contain a number of individual water
bodies that are not acid sensitive.  We also recognize that there is no basis for independently
evaluating the degree  of protectiveness afforded by any specific percentile value, since it is the
combination of form and level, in conjunction with the  indicator and averaging time, which
determine the degree of protectiveness. In  light of this, we conclude that it is appropriate to
consider initially a range of percentile values, from well above the 50th percentile, or median, of
the distribution to somewhat below the highest value. For purposes of this policy assessment,  we

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have considered percentile values in the range of the 70th to the 90th percentile.  We conclude that
it would not be appropriate to represent an ecoregion with the highest or near highest critical load
so as to avoid potential extreme outliers that can be seen to exist at the extreme end of the data
distributions, which would not be representative of the population of acid sensitive water bodies
within the ecoregion.  Also, in considering ecoregions that are inherently acid sensitive, we have
limited the lower end of our range of consideration to the 70th percentile, a value well above the
median of the distribution.
       With regard to relatively non-acid sensitive ecoregions, we recognize that while such
ecoregions are generally less sensitive to acidifying deposition from oxides of nitrogen and
sulfur, they may contain a number of water bodies that are acid sensitive  This  category includes
ecoregions that are well protected from acidification effects due to natural production of base
cations and high ANC levels, as well as naturally acidic systems with limited base cation
production and consequently very low  critical loads. Therefore, the use of a critical load  that
would be associated with highly sensitive water bodies in a naturally acidic system would
impose  a high degree of relative protection in terms of allowable ambient air concentrations of
oxides of nitrogen and sulfur and related deposition, while potentially affording little or no
public welfare benefit from attempting to improve a naturally acidic system.  Based on these
considerations, we conclude it is appropriate to consider the use of a range of percentile values
that extends lower than the range identified above for acid sensitive ecoregions. Consideration
of a lower percentile would avoid representing a relatively non-acid sensitive ecoregion by  a
critical load associated with relatively more sensitive water bodies.  In particular, we conclude it
is appropriate to focus on  the median or 50th percentile of the distribution of critical loads so as
to avoid over-protection in such ecoregions.  Recognizing that relatively non-acid sensitive areas
generally are not sampled to the extent that acid sensitive regions are, it also is appropriate to
consider using the median critical load of all relatively non-acid sensitive areas.

•  How do we calculate AAI factors in data-limited ecoregions?
       The initial delineation of acid sensitive and relatively non-acid sensitive ecoregions was
based on available ANC and alkalinity data.  Areas not meeting the ANC criteria described
above are categorized as relatively non-acid sensitive.  The development of a reasonable
distribution of critical loads  for water bodies within an ecoregion for the purpose of identifying
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the nth percentile representative critical load requires additional data, including more specific
water quality data for major cations and anions. This means that the water bodies that can be
used to develop a distribution of critical loads is generally a subset of those water bodies for
which ANC data are available  Consequently, there are certain ecoregions with sparse data that
are not suitable for developing a distribution of critical loads.
       As noted above, we judge that it is not appropriate to develop such distributions based on
data from less than ten water bodies within an ecoregion.  Such ecoregions, which included only
relatively non-acid sensitive ecoregions, were characterized as being data-limited. We identified
12 such ecoregions, and for these ecoregions we considered alternative approaches to specifying
values for the terms CLr and Qr for the purpose of determining values for each of the factors in
the AAI equation.  For these data-limited ecoregions, we judge that it is appropriate to use the
median values of CLr and Qr from the distributions of these terms for all other relatively non-acid
sensitive ecoregions, rather than attempting to use severely limited data to develop a value for
these terms based solely on data from such an ecoregion.  We note that this data limitation is not
a concern in specifying values for the other terms in the AAI equation for such ecoregions, since
those terms are based on data from the 2005 CMAQ model  simulation, which covers all
ecoregions across the contiguous United States.
Data coverage for Hawaii. Alaska,  and the U.S. Territories
       The above methods apply to those ecoregions within the contiguous U.S.  For those areas
outside the continental U.S., there is currently a lack of available data to characterize the
sensitivity of such areas,  as well as a  lack of water body-specific data and CMAQ-type modeling
to specify values for the terms Fl through F4 in the AAI equation.  Thus, we have considered
possible alternative approaches to specifying values for factors Fl through F4 in the AAI
equation for these areas.
       One such approach could be to specify area-specific values for the factors based on
values derived for ecoregions with  similar acid sensitivities, to the extent that relevant
information can be obtained to determine such similarities.  Such an approach would involve
conducting an analysis to characterize similarities in relevant ecological attributes between
ecoregions in the contiguous U.S. and these areas outside the contiguous U.S. so as  to determine
the appropriateness of utilizing ecoregion-specific values for the CLr and Qr terms from one or
more ecoregions within the contiguous U.S.  This approach would also involve conducting
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additional air quality modeling for these area that are outside the geographical scope of the
currently available CMAQ model simulations, so as to develop the other information necessary
to specify values for factors F2 through F4 for these areas.
       A second approach could rely on future data collection efforts to establish relevant
ecological data within these areas that, together with additional air quality modeling, could be
used to specify area-specific values for factors Fl through F4.  Until such time as relevant data
become available, these areas could be treated the same as data-limited ecoregions in the
contiguous U.S. that are relatively non-acid sensitive.
       Staff concludes that either approach would introduce substantial uncertainties that arise
from attempting to extrapolate values based on similarity assumptions or arbitrarily assigning
values for factors in the AAI equation that would be applicable to these areas outside the
contiguous U.S. In light of such uncertainties, we conclude that it would also be appropriate to
consider relying on the existing NC>2 and 862 secondary standards in these areas for protection of
any potential direct or deposition-related ecological effects that may be associated with the
presence of oxides of nitrogen and sulfur in the ambient air.  In staffs view, relying on existing
secondary standards in these areas is preferable to using a highly uncertain approach to allow for
the application of a new standard based on the AAI in the absence of relevant area-specific data.

7.3    AVERAGING TIME
       Aquatic acidification can occur over both long- and short-term timescales. Long-term
cumulative deposition of nitrogen and sulfur is reflected in the chronic acid-base balance of
surface waters as indicated by measured annual ANC levels. Similarly, the use of steady  state
critical load modeling, which generates critical loads in terms of annual cumulative deposition  of
nitrogen and sulfur, means that the focus of ecological effects studies based on critical loads is
on the long-term equilibrium status of water quality in aquatic  ecosystems.  Much of the
evidence of adverse ecological effects associated with aquatic acidification, as discussed above
in chapter 3, is associated  with chronically low ANC levels. Protection against a chronic ANC
level that is too low is provided by reducing overall annual average deposition levels for nitrogen
and sulfur.
       Reflecting this focus on long-term acidifying deposition, we developed the AAI that links
ambient air indicators to deposition-related ecological effects, in terms of several factors,  Fl
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through F4. As discussed above, these factors are all calculated as annual average values,
whether based on water quality and hydrology data or on CMAQ model simulations. In the
context of a standard defined in terms of the AAI, staff concludes that it is appropriate to
consider the same annual averaging time for the ambient air indicators as is used for the factors
in the AAI equation.
       We also recognize that short-term (i.e., hours or days) episodic changes in water
chemistry, often due to changes in the hydrologic flow paths (Chen et al. 1984), can have
important biological effects in aquatic ecosystems. Such short-term changes in water chemistry
are termed "episodic acidification." Some streams may have chronic or base flow chemistry that
is generally healthy for aquatic biota, but may be subject to occasional acidic episodes with
potentially lethal consequences. Thus, short-term episodic ecological effects can occur even in
the absence of long-term chronic acidification effects.
       Episodic declines in pH and ANC are nearly ubiquitous in drainage waters throughout the
eastern United States. Episodic acidification can result from several mechanisms related to
changes in hydrologic flow paths. For example, snow can store nitrogen deposited throughout
the winter and snowmelt can then release this stored nitrogen, together with  nitrogen derived
from nitrification in the soil itself, in a pulse that leads to episodic acidification in the absence of
increased  deposition during the actual episodic acidification event.  We note that inputs of
nitrogen and sulfur from snowpack and atmospheric deposition largely cycle through soil. As a
result, short-term direct deposition  inputs are not necessarily important in episodic acidification.
Thus, as noted in chapter 3 of the ISA, protection against episodic acidity events can be achieved
by establishing a higher chronic ANC level.
       Taken together, the above considerations support the conclusion that it is appropriate to
consider the use of a long-term average for the ambient air indicators NOy and SOx for an
aquatic acidification standard defined  in terms of the AAI. The use of an annual averaging time
for NOy and SOx concentrations would be appropriate to provide protection against low chronic
ANC levels, which in turn would protect against both long-term acidification and acute acidic
episodes.
       We have also considered interannual variability in both ambient air quality and in
precipitation, which is directly related to the deposition of oxides of nitrogen and sulfur from the
ambient air. While ambient air concentrations show year-to-year variability, we note that often

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the year-to-year variability in precipitation is considerably greater, given the highly stochastic
nature of precipitation. The use of multiple years over which annual averages are determined
would dampen the effects of interannual variability in both air quality and precipitation.  For the
ambient air indicators, the use of multiple-year averages would also add stability to calculations
used to judge whether an area meets a standard defined in terms of the AAI.  Consequently, staff
concludes that an annual averaging time based on the average of each year over a consecutive 3
to 5 year period is appropriate to consider for the ambient air indicators NOy and SOx.  In
reaching this conclusion,  we note that in its comments  on the second draft PA,  CAS AC agreed
that a 3 to 5  year averaging time was appropriate to consider (Russell and Samet, 201 Ob; p.4).

7.4    LEVEL
       As discussed above in section 7.2.1, ANC is the ecological indicator best suited to reflect
the sensitivity of aquatic ecosystems to acidifying deposition from oxides of nitrogen and sulfur
in the ambient air. ANC  is an indicator of the aquatic acidification expected to occur given the
natural buffering capacity of an ecosystem and the loadings of nitrogen and sulfur resulting from
atmospheric deposition. Thus, in this PA we have developed a new standard for aquatic
acidification that is based on the use of chronic ANC as the ecological indicator as a component
in the AAI.
       The level of the standard would be defined in terms of a single, national value of the
AAI.  The standard would be met at a monitoring site when the multi-year average of the annual
values of the AAI was equal to or above the specified level of the standard,8  where the annual
values of the AAI would be calculated based on the AAI  equation using the assigned ecoregion-
specific values for factors Fl through F4 and monitored annual average NOy and SOx
concentrations. Since the AAI equation is based on chronic ANC as the ecological  indicator, the
level chosen for the standard would reflect a target chronic ANC value. The assigned F factors
for each ecoregion would be determined by EPA based on water quality and  hydrology data,
CMAQ modeling, the selected percentile value that is used to identify a representative critical
load within the ecoregion, and the level of the standard. The combination of the form of the
standard (section 7.2), defined by the AAI equation and the assigned values of the F factors in
 Unlike other NAAQS, where the standard is met when the relevant value is at or below the level of the standard
since a lower standard level is more protective, in this case a higher standard level is more protective.
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the equation, other elements of the standard including the ambient air indicators (section 7.1) and
their averaging time (section 7.3), and the level of the standard determines the allowable levels
of NOy and SOx in the ambient air within each ecoregion.  All of the elements of the standard
together determine the degree of protection from adverse aquatic acidification effects associated
with oxides of nitrogen and sulfur in the ambient air.  The level of the standard plays a central
role in determining the degree of protection provided and is discussed below.
       We focus primarily on information that relates degrees of biological impairment
associated with adverse ecological effects to aquatic ecosystems to alternative levels of ANC in
reaching staff conclusions regarding the range of target ANC levels that is appropriate  to
consider for the level of the standard. We develop the rationale for identifying a range of target
ANC levels that is appropriate to consider by addressing questions related to the following areas:
   (1) associations between ANC and pH levels to provide an initial bounding for the range of
       ANC values to be considered;
   (2) evidence that allows for the delineation of specific ANC ranges associated with varying
       degrees of severity of biological impairment ecological effects;
   (3) the  role of ANC in affording protection against episodic acidity;
   (4) implications of the time lag response of ANC to  changes in deposition;
   (5) past and current examples of target ANC values applied in environmental management
       practices; and
   (6) data linking public welfare benefits and ANC.

•  What range of pH levels is useful to help inform an initial bounding of target ANC
   levels?
       As discussed above in chapter 3, specific levels of ANC are associated with differing
levels of risk of biological impairment in aquatic ecosystems, with higher levels of ANC
resulting in lower risk of ecosystem impacts, and lower ANC levels resulting in risk of both
higher intensity of impacts and a broader set of impacts. While ANC is not the causal  agent
determining biological  effects in aquatic ecosystem, as discussed above in section 7.2,  it is a
useful metric for determining the level at which a water  body is protected against risks of
acidification. There is a direct correlation between ANC and pH levels which, along with
dissolved aluminum,  are more closely linked to the biological causes of ecosystem response to
acidification.
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       Because there is a direct correlation between ANC and pH levels, we can inform the
selection of target ANC levels in part through information on effects of pH as well as direct
studies of effects related to ANC. Levels of pH are closely associated with ANC in the pH range
of approximately 4.5 to 7 (Figure 7-2). Within this range, higher ANC levels are associated with
higher pH levels.  At a pH level of approximately 4.5, further reductions in ANC generally do
not correlate with pH, as pH levels remain at approximately 4.5 while ANC values fall
substantially. Similarly, at a pH value of approximately 7, ANC values can continue to increase
with no corresponding increase in pH.  As pH is the primary causal indicator of aquatic
acidification related effects, this suggests that ANC values below approximately -50 ueq/L (the
apparent  point in the relationship  between pH and ANC where pH reaches a minimum) are not
likely to result in further damage, while ANC values around and above approximately 100 ueq/L
(the apparent region in the relationship where pH reaches a maximum) are not likely to confer
additional protection. As a result, our initial focus is on target ANC values in the range of-50 to
100 ueq/L.
•   What specific ANC ranges are related to varying degrees of effects on aquatic
    ecosystems?
       As discussed above in chapter 3 and section 7.2, the  number offish species present in a
water body has been shown to be positively correlated with the ANC level in the water, with
higher values supporting a greater richness and diversity offish species (Figures 7-3 and 7-4).
The diversity and distribution of phyto-zooplankton communities also are positively correlated
with ANC.
       Within the ANC range from approximately -50 to 100 ueq/L, linear and sigmoidal
relationships are observed (shown in Figures 7-3 and 7-4, respectively) between ANC and
ecosystem effects.  On average, fish species richness is  lower by one fish species for every 21
ueq/L decrease in ANC in Shenandoah National Park streams (ISA, section 3.2.3.4).  As shown
in Table 3-3, ANC levels  have been grouped into five categories related to expected ecological
effects, including categories of acute concern ( <0 ueq/L), severe concern (0-20 ueq/L), elevated
concern (20-50 ueq/L), moderate concern (50-100 ueq/L), and low concern (>100 ueq/L).   This
categorization is supported by a large body of research completed throughout the eastern United
States (Sullivan et al, 2006).
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       Water bodies with ANC values less than or equal to 0 ueq/L are chronically acidic. Such
ANC levels can lead to complete loss of species and major changes in the ability of water bodies
to support diverse biota, especially in water bodies that are highly sensitive to episodic
acidification. Based on the above considerations, staff has focused on target ANC levels no
lower than 0 ueq/L.
       Biota generally are not harmed when ANC values are >100 ueq/L, due to the low
probability that pH levels will be below 7.  In the Adirondacks, the number offish species also
peaks at ANC values >100 ueq/L. This suggests that at ANC levels greater than 100 ueq/L, little
risk from acidification exists in many aquatic ecosystems.  At ANC levels below 100 ueq/L,
overall health of aquatic communities can be maintained, although fish fitness and community
diversity begin to decline.  At ANC levels ranging from 100 down to 50 ueq/L, there is
increasing likelihood that the fitness of sensitive species (e.g., brook trout, zooplankton) will
begin to decline. When ANC concentrations are below 50 ueq/L, the probability of acidification
increases substantially, and negative effects on aquatic biota are observed, including large
reductions in diversity offish species and changes in the health offish populations, affecting
reproductive ability and fitness.  We recognize that while there is evidence that ANC levels
above 50 can confer additional protection from adverse ecological effects associated with aquatic
acidification in some sensitive ecosystems, the expectation that such incremental protection from
adverse effects will continue up to an ANC level of 100 is substantially reduced. In staffs view,
the above considerations support a focus on target ANC levels up to a level greater than 50 ueq/L
but below 100 ueq/L, such as up to a level of 75  ueq/L.
       In considering the available scientific evidence, as summarized here and discussed in
more detail in the ISA and REA, we note that in its review of the second draft PA CAS AC
expressed the following views about the range of biological responses that corresponds to this
range of ANC levels (i.e., 0-100 ueq/L):
       "There will likely be  biological effects of acidification at higher ANC values
       within this range, and there are relatively insensitive organisms that are not
       impacted at ANC values at the low end of this  range. Adverse effects of
       acidification on aquatic biota are fairly certain  at the low end of this range of
       ANC and incremental benefits of shifting  waters to higher ANC become more
       uncertain at higher ANC levels.  There is substantial confidence that there are
       adverse effects at ANC levels below 20 ueq/L, and reasonable confidence that
       there are adverse effects below 50 ueq/L.  Levels of 50 ueq/L and higher would
       provide additional protection, but the Panel has less confidence in the significance
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       of the incremental benefits as the level increases above 50 ueq/L." (Russell and
       Samet, 201 Ob; pp. 15-16)
       The above considerations, including the views of CAS AC, provide support for focusing
on target ANC levels in the range of 20 to 75 ueq/L.
•  What is the role of ANC in protecting against low pH levels and episodic acidity?
       Across the broad range of ANC values from 0 to 100 ueq/L, ANC affords protection
against the likelihood of decreased pH (and associated increases in Al).  In general, the higher
the ANC within this range, the lower the probability of reaching low pH levels where direct
effects such as increased fish mortality (Table 3-1) occur.  Accordingly, greater protection
would be achieved by target chronic ANC values set high enough to avoid pH depression to
levels associated with elevated risk.
       The specific relationship between ANC and the probability of reaching pH levels of
elevated risk varies by water body and fish species. ANC levels below 20  ueq/L are generally
associated with high probability of low pH, leading to death or loss of fitness of biota that are
sensitive to acidification (ISA, section 5.2.2.1; REA, section 5.2.1.2).  At these levels, during
episodes of high acidifying deposition, brook trout populations may experience lethal effects. In
addition, the diversity and distribution of zooplankton communities decline sharply at ANC
levels below 20 ueq/L. Overall, there is little uncertainty that significant effects on aquatic biota
are occurring at ANC levels below 20 ueq/L.
       It is clear that at ANC levels approaching 0 ueq/L (Table 3-1),  there is significant
impairment of sensitive aquatic ecosystems with almost complete loss offish species. Avoiding
ANC levels  approaching 0 ueq/L is particularly relevant to episodic spikes in acidity that occur
during periods of rapid snow melt and during and after major precipitation events. Since the
ANC range we are discussing here reflects average, long-term sustained values, consideration
should be given to protecting against episodic drops in ANC values to a  level as low as 0 ueq/L.
The above considerations do not provide support for a target chronic ANC level as low as 0
ueq/L for a standard that would protect against significant harm to aquatic  ecosystems, including
harm from episodic acidification. In staffs view, these  considerations also support a lower end
of the range for consideration no lower than 20 ueq/L.
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       We note that CASAC agreed with this view in its comments on the second draft PA
(Russell and Sam et, 201 Ob; p.  16).  CA SAC noted that "there are clear and marked
biological effects at AN C values near 0 ueq/L, so this is probably not an appropriate target
value" for the AAI. With regard to the likelihood of impairment of aquatic ecosystems due to
episodic acidification, in terms of specific target levels for chronic ANC, CAS AC expressed the
following view:
       "Based on surface waters  studied in the Northeast, decreases in ANC associated
       with snowmelt [are] approximately 50 ueq/L. Thus, based on these studies, a
       long term ANC target level of 75 ueq/L would generally guard against effects
       from episodic acidification down to a level of about 25 ueq/L." (Russell and
       Samet, 201 Ob; p. 26)
•  Wiat arethe implications of considering eoosystem respcnsetime?
       When considering a standard level to  protect against aquatic acidification, it is
appropriate to take into account both the time period to recovery as well as the potential for
recovery in acid sensitive ecoregions.  Ecosystems become adversely impacted by acidifying
deposition over long periods of time and have variable time frames and abilities to recover
from such perturbations Modeling presented  in the REA (REA, section 4.2.4) showsthe
estimated ANC valuesfor Adirondack lakes and Shenandoah streamsunder pre-
acidification conditions and indicates that for a small percentage of lakes and streams,
natural ANC levels would have been below 50 ueq/L.  Therefore, for these  water bodies,
reductions in acidifying deposition are not  likely to achieve an ANC of 50 ueq/L or
greater.  Conversely, for some lakes and streams the level of perturbation from long
periods of acidifying deposition has resulted in very low ANC values com pared to
estimated natural conditions For such water bodies, the time to recovery would be largely
dependent on future inputs of acidifying deposition.
       Setting a standard level in terms of  a target chronic ANC level is based on the long-
term response of aquatic ecosystems  The time required for a water body to achieve the
target ANC level given a decrease in ambient  air concentrations of NOy and SOx and
related acidifying deposition such  that  the  critical load for that target ANC is not exceeded
is often decades if not centuries  In recognition of the potential public welfare benefits of
achieving the target AN C in a shorter  time frame, the concept of target  loads had been
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developed.  Target loads represent the depositional loading that is expected to achieve a
particular level of the ecological indicator by a given time.  For example, to achieve an
ANC level of 20 u^q/L by 2030, it might be necessary to specify a higher target ANC
level of, for example, 50 neq/L, such that the depositional loading would be reduced more
quickly than would occur if the depositional loading was based on achieving a target ANC
level of 20 [^eq/L as a long-term equilibrium level.  In this example, the target ANC of 50
^eq/L would ultimately be realized many years later.
      The above considerations have implications for selecting an appropriate standard
level, in that the standard level affects not only the ultimate degree of protection that
would be afforded by the standard, but also the time frame in which such protection would
be realized.  However, we recognize that there is a great deal of heterogeneity in response
times among water bodies and that there is only very limited information from dynamic
modeling that would help to quantify recovery time frames in areas across the country.  As
a consequence,  we recognize that quantification of a general relationship between critical
loads associated with a specific long-term target ANC level and target loads associated with
achieving the target ANC level within a specific time frame is not currently possible.  Thus,
while the time frame for recovery is an important consideration in selecting an appropriate
range of  levels to consider, we conclude that it can only be considered in a qualitative
sense at this time.
•  Wiat ANCtarget levels have been set by other organizations to protect against aquatic
   acidification?
      A number of  regional organizations, states, and international organizations have
developed critical load frameworksto protect against acidification of sensitive aquatic
ecosystems In considering the appropriate range of target ANC levelsfor consideration in
this review,  it is informative to evaluate the target ANC levels selected by these different
organizations, as well as the rationale provided in support of the selected levels  Chapter 4
provides a detailed discussion of how critical loads have been developed and used in other
contexts  This section summarizes such specific target values and their rationales
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       The UNECE has developed critical loads in support of international emissions
reduction agreements  As noted in chapter 4, critical loads were established to protect 95
percent of surface waters in  Europe from an ANC less than 20 ueq/L based on protection
of brown trout.  Individual countries have set alternative ANC targets; for example,
Norway targets an ANC of 30 ueq/L based on protection of Atlantic salmon.
       Several states have established target ANC or pH values related to protection of
lakes and streamsfrom acidification.  While recognizing that some lakes in theAdirondacks
will have a naturally low pH, the state of New York has established a target pH value of
6.5 for lakesthat are not naturally below 6.5.  As noted above, this level  is associated
with an ANC value that is likely to be between  20 and 50 ueq/L or possibly higher.  New
Hampshire and Vermont have set ANC targets of 60 ueq/L and 50 ueq/L,  respectively.
Tennessee has established site-specific target ANC values based on assessments of natural
acidity, with a default value of 50 ueq/L when specific data are not available.
       Taken together, these policy responsesto concerns about ecological effects
associated with aquatic acidification indicate that target ANC values between 20 and 60
ueq/L have been selected by states and other nationsto provide protection of lakes and
streams in some of the more sensitive aquatic ecosystems
•  What relevant information is available that links public welfare benefits to alternative
   target ANC levels?
       The point at which effects on public welfare become adverse is not defined in the CAA.
Characterizing a known or anticipated adverse effect to public welfare is an important
component of developing any secondary NAAQS. According to the CAA, welfare effects
include:
       "...effects on soils, water, crops, vegetation, manmade materials, animals,
       wildlife, weather, visibility, and climate, damage to and deterioration of property,
       and hazards to transportation, as well as effect on economic values and on
       personal comfort and well-being, whether caused by transformation, conversion,
       or combination with other air pollutants." (CAA, section 302(h)).
While the text above lists a number of welfare effects, the NAAQS is  aimed at protection
from adverse effects to public welfare.  Consideration of adversity to public welfare in
the context of the secondary NAAQS for oxides of nitrogen and sulfur can be informed
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by information about losses in ecosystem services associated with acidifying deposition
and the potential economic value of those losses, as discussed above in chapter 4.
       Ecosystem service losses at alternative ANC levels are difficult to enumerate. However,
in general there are categories of ecosystem services, discussed in chapter 4, that are related to
the specific ecosystem damages expected to occur at alternative ANC levels. Losses in fish
populations due to very low ANC (below 20 ueq/L) are likely associated with significant losses
in value for recreational and subsistence fishers. Many acid sensitive lakes are located in areas
with high levels of recreational fishing activity.  For example, in the northeastern U.S., where
nearly 8 percent of lakes are considered acidic, more than 9 percent of adults participate in
freshwater fishing, with an estimated value of approximately $5 billion in 2006. This suggests
that improvements in lake fish populations are likely associated with significant recreational
fishing value.
       Inland surface waters also  provide cultural services such as aesthetic and existence value
and educational services.  To the extent that piscivorous birds and other wildlife are harmed by
the absence offish in these waters, hunting and birdwatching activities  are likely to be adversely
affected. A case study of the value to New York residents of improving the health of lakes in the
Adirondacks found significant willingness to pay for those improvements.   When scaled to
evaluate the improvement in lake health from achieving ANC values of 20 to 50 ueq/L, the
study implies benefits to the New  York population roughly on the order of $600 million per year
(in constant 2007$).  The survey administered in this study recognized that  participants were
thinking about the full range of services provided by the lakes in question - not just the
recreational fishing services.  Therefore the estimates of willingness to  pay  include resident's
benefits for potential hunting and birdwatching activities and other ancillary services. These
results are just for New York populations.  If similar benefits exist for improvements in other
acid sensitive lakes, the economic value to U.S.  populations could be very substantial,  suggesting
that, at least by one measure of impact on public welfare, impacts associated with ANC less than
50 ueq/L may be adverse to public welfare.
•  What are staff conclusions with regard to a range of standard levels that is appropriate
   to consider to protect against deposition-related aquatic acidification effects associated
   with oxides of nitrogen and sulfur in the ambient air?
       Based on all the above considerations, staff concludes that consideration should be given
to a range of standard levels from  20 to 75 ueq/L. The available evidence indicates that target
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ANC levels below 20 ueq/L would be inadequate to protect against substantial ecological effects
and potential catastrophic loss of ecosystem function in some sensitive aquatic ecosystems.
While ecological effects occur at ANC levels below 50 ueq/L in some sensitive ecosystems, the
degree and nature of those effects are less significant than at levels below 20 ueq/L. Levels at
and above 50 ueq/L would be expected to provide additional protection, although uncertainties
regarding the potential for additional protection from  adverse ecological effects are much larger
for target ANC levels above about 75 ueq/L as effects are generally appreciably less sensitive to
changes in ANC at such higher levels.
       In reaching this conclusion, staff took into consideration the extent to which a target
ANC level within this range would protect against episodic as well as long-term ecological
effects.  Levels in the mid- to upper part of this range would be expected to provide greater
protection against short-term, episodic peaks in aquatic acidification, while lower levels within
this range would give more weight to protection from long-term rather than episodic
acidification.  Similarly, levels in the mid- to upper part of this range would be expected to result
in shorter time periods for recovery given the lag in ecosystem response in some sensitive
ecosystems relative to levels in the lower part of this range. We also note that this range
encompasses target ANC values that have been established by various States and regional and
international organizations to protect against acidification of aquatic ecosystems.
       We recognize that the level of standard together with the other elements of the standard,
including the ambient air indicators, averaging time, and form, determine the overall
protectiveness of the standard. Thus, consideration of a standard level should reflect the
strengths and limitations of the evidence and assessments as well as the inherent uncertainties in
the development of each of the elements of the standard. The implications of considering
alternative  standards, defined in terms of alternative combinations of levels and percentile values
that are a critical component of the form of the standard, are discussed below in section 7.5. Key
uncertainties in the various  components of the standard are summarized and considered below in
section 7.6.
7.5    Considerations associated with alternative standards
       To provide some perspective on the implications of various alternative standards, staff
assessed the number of acid sensitive ecoregions that would likely not meet a set of alternative
standards.  The alternative standards considered in this assessment were based on combinations
                                           7-53

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of alternative levels, within the range of 20 to 75 jieq/L identified above in section 7.4, and
alternative forms, characterized by alternative representative percentile values within the range
of the 70th to 90th percentile identified above in section 7.2.5.  These alternative standards are
also defined in terms of the other elements of the standard:  ambient air indicators NOy and SOx
identified above in section 7.1; other elements of the form of the standard, including ecoregion-
specific values for factors Fl through F4 in the AAI equation, specified as discussed above in
section 7.2; and an annual averaging time for NOy and SOx, as discussed above in section 7.3.
With regard to the averaging time,  we did not consider multi-year averaging of the calculated
annual AAI values due to data limitations, including, for example, the lack of CMAQ modeling
for  multiple consecutive years. In this assessment, we characterize an ecoregion as likely not
meeting a given alternative standard if the calculated AAI value is less than the level of the
standard, recognizing that higher AAI values are more protective than lower values.
       The results of this assessment are presented below for  each of the 29 ecoregions
characterized as acid sensitive. Calculated annual AAI values, which in essence are "design
values" at the  ecoregion level,  are shown below in Table 7-1 for each acid sensitive ecoregion
for  each alternative  standard considered. Based on these AAI values, Table 7-2 then summarizes
the  number of acid sensitive ecoregions that  would likely not meet each of the alternative
standards considered. We also calculated AAI for all ecoregions categorized as relatively non-
acid sensitive, as shown in Table D-5 in Appendix D. In all cases, these ecoregions were likely
to meet all of the alternative standards considered in this assessment.
       As described above in section 7.2, the AAI values presented here are based in part on
data from 2005 CMAQ model simulations, which was used to generate values for F2 through F4
in the AAI equation as well as to estimate ambient air concentrations of NOy and SOx that
reflect recent air quality in the absence of currently available monitored concentrations in
sensitive ecoregions across the country. Water quality and hydrology data from water bodies
within each ecoregion were also used in calculating the AAI values. Such data were initially
used to calculate critical loads  for each water body with sufficient data within an ecoregion so as
to identify the nth percentile critical load representative of the  ecoregion used in calculating the
Fl factor for the ecoregion. In developing the distribution of critical loads for each ecoregion,
three approaches were considered to define the water bodies included in the distribution.  These
approaches included using (1)  all water bodies with available  data, (2) screening out water

                                           7-54

-------
bodies with SO4"2 levels > 400 jieq/L as an indicator acid mine drainage activities and (3) adding
additional screening to eliminate water bodies with DOC values >10mg/L or critical loads less
than 10 meq/m2-yr, as indicators of naturally acidic systems.  Here, we present results of
analyses that did not include any screens as there were only marginal differences in the results
with or without applying the screens. The inclusion of all water bodies also provided useful
diagnostic information.   Because the representative critical load for an ecoregion is an important
quantity that reflects the nth percentile sensitivity, consideration should be given to alternative
methods such as using an interpolated nth percentile value based on the distribution of critical
loads rather than using a water body-specific critical load based on its rank order within the
distribution.
                                            7-55

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Table 7- la. Calculated AAI values for acid sensitive ecoregions across the range of nth percentiles for
an alternative level of 20 jieq/L. (highlighted values indicate regions not likely to meet
an alternative standard)

6.2.4
6.2.3
6.2.7
8.5.4
5.3.1
6.2.10
8.1.3
8.1.7
5.3.3
8.1.8
6.2.5
5.2.1
6.2.15
8.4.1
8.4.2
8.4.3
6.2.13
8.5.1
6.2.12
6.2.14
8.4.4
8.3.5
8.3.4
8.4.9
8.4.6
8.4.7
8.5.3
8.4.8
8.3.7

Canadian Rockies
Columbia Mountains/Northern Rockies
Cascades
Atlantic Coastal Pine Barrens
Northern Appalachian and Atlantic Maritime
Highlands
Middle Rockies
Northern Appalachian Plateau and Uplands
Northeastern Coastal Zone
North Central Appalachians
Maine/New Brunswick Plains and Hills
North Cascades
Northern Lakes and Forests
Idaho Batholith
Ridge and Valley
Central Appalachians
Western Allegheny Plateau
Wasatch and Uinta Mountains
Middle Atlantic Coastal Plain
Sierra Nevada
Southern Rockies
Blue Ridge
Southeastern Plains
Piedmont
Southwestern Appalachians
Boston Mountains
Arkansas Valley
Southern Coastal Plain
Ouachita Mountains
South Central Plains
70th
933.4
353.5
90.2
-154.6
58.3
180.0
227.0
42.0
-60.8
89.7
138.4
51.4
66.8
-72.3
-78.3
412.4
297.8
-17.2
49.1
120.6
-65.5
-51.2
131.6
35.3
65.1
90.1
-31.2
89.3
287.1
75th
740.3
267.3
72.2
-172.7
49.0
122.1
173.4
22.6
-74.4
84.4
130.7
38.7
62.0
-95.1
-109.0
280.7
255.3
-29.5
38.2
98.5
-73.5
-59.2
102.7
-18.5
65.1
82.5
-61.8
74.6
279.6
80th
685.6
190.3
46.2
-174.6
33.6
99.0
165.7
9.3
-87.4
71.0
112.9
25.9
59.3
-117.6
-147.0
47.4
230.6
-64.0
28.1
85.7
-83.3
-73.1
72.7
-29.4
27.7
66.2
-105.1
64.6
213.3
85th
551.0
136.5
31.3
-182.4
21.3
81.6
120.4
-4.4
-97.8
65.5
93.8
14.1
48.0
-143.7
-169.5
-20.8
174.6
-131.6
22.2
67.6
-93.1
-91.0
45.7
-69.8
8.7
50.7
-143.2
51.3
136.4
90th
84.5
106.3
19.1
-193.6
6.4
69.4
85.6
-23.6
-112.5
48.5
65.8
3.8
41.6
-177.8
-186.2
-97.9
136.6
-169.4
12.6
50.8
-104.9
-106.7
11.8
-121.5
-24.4
-1.0
-154.9
-3.3
47.3
7-56

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Table 7-lb. Calculated AAI values for acid sensitive ecoregions across the range of nth percentiles for
an alternative level of 35 jieq/L. (highlighted values indicate regions not likely to meet
an alternative standard)

6.2.4
6.2.3
6.2.7
8.5.4
5.3.1
6.2.10
8.1.3
8.1.7
5.3.3
8.1.8
6.2.5
5.2.1
6.2.15
8.4.1
8.4.2
8.4.3
6.2.13
8.5.1
6.2.12
6.2.14
8.4.4
8.3.5
8.3.4
8.4.9
8.4.6
8.4.7
8.5.3
8.4.8
8.3.7

Canadian Rockies
Columbia Mountains/Northern Rockies
Cascades
Atlantic Coastal Pine Barrens
Northern Appalachian and Atlantic Maritime
Highlands
Middle Rockies
Northern Appalachian Plateau and Uplands
Northeastern Coastal Zone
North Central Appalachians
Maine/New Brunswick Plains and Hills
North Cascades
Northern Lakes and Forests
Idaho Batholith
Ridge and Valley
Central Appalachians
Western Allegheny Plateau
Wasatch and Uinta Mountains
Middle Atlantic Coastal Plain
Sierra Nevada
Southern Rockies
Blue Ridge
Southeastern Plains
Piedmont
Southwestern Appalachians
Boston Mountains
Arkansas Valley
Southern Coastal Plain
Ouachita Mountains
South Central Plains
70%
934.4
342.8
92.1
-160.7
58.8
179.1
225.5
42.3
-59.8
89.5
144.4
52.7
60.2
-72.3
-80.3
415.2
287.3
-16.2
47.4
120.2
-65.4
-55.5
131.2
31.1
65.1
89.3
-29.1
89.5
291.9
75%
736.5
257.3
74.7
-173.7
46.2
121.7
173.7
23.2
-71.8
84.6
125.1
39.6
56.7
-94.0
-107.3
281.4
243.0
-32.0
40.5
102.3
-71.6
-63.4
96.1
-12.2
65.1
85.7
-61.7
78.9
275.4
80%
692.4
188.0
47.8
-174.3
32.1
99.2
163.9
8.9
-88.7
72.1
110.3
25.7
53.7
-116.3
-143.2
46.2
221.8
-61.4
30.6
87.3
-82.4
-72.1
72.6
-29.1
28.9
71.4
-106.5
67.4
210.4
85%
556.9
122.5
30.7
-180.7
20.1
81.0
121.4
-6.1
-95.3
65.9
86.5
14.3
49.6
-144.8
-173.0
-22.1
184.4
-133.2
21.8
66.4
-93.6
-96.2
43.7
-71.4
8.7
50.6
-138.6
49.3
133.1
90%
80.9
106.8
21.4
-192.9
3.4
63.3
87.3
-24.2
-114.3
47.1
71.6
1.0
39.8
-175.4
-182.0
-94.1
126.3
-163.4
13.8
53.1
-104.6
-107.9
14.0
-121.3
-21.2
-2.0
-150.8
-4.8
47.3
7-57

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Table 7-lc. Calculated AAI values for acid sensitive ecoregions across the range of nth percentiles for
an alternative level of 50 jieq/L. (highlighted values indicate regions not likely to meet
an alternative standard)

6.2.4
6.2.3
6.2.7
8.5.4
5.3.1
6.2.10
8.1.3
8.1.7
5.3.3
8.1.8
6.2.5
5.2.1
6.2.15
8.4.1
8.4.2
8.4.3
6.2.13
8.5.1
6.2.12
6.2.14
8.4.4
8.3.5
8.3.4
8.4.9
8.4.6
8.4.7
8.5.3
8.4.8
8.3.7

Canadian Rockies
Columbia Mountains/Northern Rockies
Cascades
Atlantic Coastal Pine Barrens
Northern Appalachian and Atlantic Maritime
Highlands
Middle Rockies
Northern Appalachian Plateau and Uplands
Northeastern Coastal Zone
North Central Appalachians
Maine/New Brunswick Plains and Hills
North Cascades
Northern Lakes and Forests
Idaho Batholith
Ridge and Valley
Central Appalachians
Western Allegheny Plateau
Wasatch and Uinta Mountains
Middle Atlantic Coastal Plain
Sierra Nevada
Southern Rockies
Blue Ridge
Southeastern Plains
Piedmont
Southwestern Appalachians
Boston Mountains
Arkansas Valley
Southern Coastal Plain
Ouachita Mountains
South Central Plains
70%
935.5
327.0
93.0
-166.7
57.0
178.2
223.9
42.1
-58.7
90.1
148.3
54.0
59.3
-75.2
-82.0
418.0
276.9
-15.2
44.9
124.1
-65.3
-57.5
125.8
26.9
65.1
88.8
-27.5
89.8
296.7
75%
732.7
262.1
72.5
-174.0
44.0
119.5
174.1
23.9
-69.2
84.8
123.8
39.8
53.6
-94.4
-105.5
282.2
230.6
-34.6
39.0
112.6
-72.9
-65.2
99.6
-6.0
65.1
88.4
-55.8
83.1
271.2
80%
699.2
192.3
51.7
-174.7
30.8
97.5
162.2
8.7
-87.5
73.4
116.8
24.7
43.9
-116.8
-140.5
43.2
199.4
-58.8
32.0
85.2
-82.7
-77.2
74.2
-28.8
30.0
76.6
-101.9
70.3
209.4
85%
562.7
132.6
29.2
-176.5
17.0
81.6
122.4
-7.2
-95.5
65.1
93.0
14.1
40.6
-144.0
-172.1
-23.4
194.2
-134.8
24.3
67.9
-92.5
-102.6
40.1
-73.1
8.7
50.5
-134.7
47.2
129.8
90%
79.4
106.4
23.9
-192.7
1.1
56.9
89.0
-24.8
-115.8
45.8
69.7
-0.2
37.5
-172.8
-182.1
-90.3
109.5
-157.4
13.7
58.1
-102.8
-112.7
14.8
-121.1
-17.9
0.7
-151.7
-6.2
47.3
7-58

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Table 7- Id. Calculated AAI values for acid sensitive ecoregions across the range of nth percentiles for
an alternative level of 75 jieq/L. (highlighted values indicate regions not likely to meet
an alternative standard).

6.2.4
6.2.3
6.2.7
8.5.4
5.3.1
6.2.10
8.1.3
8.1.7
5.3.3
8.1.8
6.2.5
5.2.1
6.2.15
8.4.1
8.4.2
8.4.3
6.2.13
8.5.1
6.2.12
6.2.14
8.4.4
8.3.5
8.3.4
8.4.9
8.4.6
8.4.7
8.5.3
8.4.8
8.3.7

Canadian Rockies
Columbia Mountains/Northern Rockies
Cascades
Atlantic Coastal Pine Barrens
Northern Appalachian and Atlantic Maritime
Highlands
Middle Rockies
Northern Appalachian Plateau and Uplands
Northeastern Coastal Zone
North Central Appalachians
Maine/New Brunswick Plains and Hills
North Cascades
Northern Lakes and Forests
Idaho Batholith
Ridge and Valley
Central Appalachians
Western Allegheny Plateau
Wasatch and Uinta Mountains
Middle Atlantic Coastal Plain
Sierra Nevada
Southern Rockies
Blue Ridge
Southeastern Plains
Piedmont
Southwestern Appalachians
Boston Mountains
Arkansas Valley
Southern Coastal Plain
Ouachita Mountains
South Central Plains
70%
937.2
316.6
89.4
-166.7
56.3
176.6
225.3
43.0
-56.9
87.9
145.6
54.9
55.7
-73.1
-81.7
422.7
259.6
-13.5
45.9
127.1
-64.8
-59.8
126.1
19.9
65.0
93.8
-31.0
90.2
304.7
75%
726.3
268.7
70.1
-173.5
43.4
125.2
174.6
25.0
-73.1
83.2
125.1
41.0
44.9
-95.5
-106.4
283.4
210.6
-38.8
39.6
112.6
-74.7
-68.9
98.6
4.4
65.0
87.0
-60.7
90.2
264.3
80%
710.5
201.8
49.9
-176.4
27.6
98.0
159.3
8.1
-85.3
71.7
116.0
26.2
38.8
-116.5
-137.9
41.2
186.2
-54.4
30.8
92.2
-83.3
-89.3
70.2
-28.4
31.9
85.3
-94.4
75.1
220.8
85%
572.5
130.5
36.2
-176.7
14.6
83.1
128.5
-5.2
-92.8
62.5
95.4
12.8
30.1
-142.2
-169.5
-25.6
168.0
-137.4
23.9
78.3
-92.9
-109.0
37.6
-75.8
8.8
50.4
-132.0
43.8
124.3
90%
63.5
40.3
11.3
-205.6
-25.3
32.7
-35.2
-47.4
-128.1
15.2
24.4
-17.4
13.6
-203.2
-263.2
-183.9
71.9
-218.4
-6.9
46.6
-116.5
-142.1
-50.6
-3741.4
-12.5
-7.4
-182.4
-44.9
20.9
7-59

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Table 7-2. Summary of the number of acid sensitive ecoregions (out of 29) not likely to meet
alternative standards based on a 2005 CMAQ simulation.
Sort by Percentile
ANC
75
50
35
20
75
50
35
20
75
50
35
20
75
50
35
20
75
50
35
20
Percentile
90
90
90
90
85
85
85
85
80
80
80
80
75
75
75
75
70
70
70
70
Number
25
22
19
19
21
19
16
13
19
16
14
10
16
13
10
9
15
11
9
8
Sort by ANC
ANC
75
75
75
75
75
50
50
50
50
50
35
35
35
35
35
20
20
20
20
20
Percentile
90
85
80
75
70
90
85
80
75
70
90
85
80
75
70
90
85
80
75
70
Number
25
21
19
16
15
22
19
16
13
11
19
16
14
10
9
19
13
10
9
8
       As expected, the number of ecoregions that likely would not meet alternative standards
increases with increasing percentile values and standard levels (Table 7-2). Out of 29 acid
sensitive ecoregions, the number of ecoregions that would likely not meet the alternative
standards considered ranges from 25 for the most protective alternative standard considered (75
|ieq/L, 90th percentile) to 8 for the least protective alternative  standard (20 |ieq/L, 70th
percentile).  It is apparent that both the percentile and the level chosen have a strong influence,
over the ranges considered, in determining the number of areas that would likely not meet this
set of alternative standards.
       In considering these results, we note first that there are two groupings of ecoregions that
would likely not meet almost all combinations of level and form (Table 7-2; Appendix D  maps
and Tables).  The first group broadly reflects Coastal Plains, including Southern Coastal Plain,
8.5.3; Southeastern Plains, 8.3.5; Middle Atlantic Coastal Plain, 8.5.1; and Atlantic Coastal Pine
                                           7-60

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Barrens, 8.3.4.  The second group is made up of southern Appalachian mountain areas, including
North Central Appalachians, 5.3.3; Ridge and Valley, 8.4.1; Central Appalachians, 8.4.2; Blue
Ridge, 8.4.4; and Southwestern Appalachians, 8.4.9.  In addition, these two groupings exhibit
the highest amounts of exceedance relative to alternative standards.
       The Northern Appalachian and Atlantic Maritime Highlands (5.3.1), which includes the
Adirondacks, and the Northern Lakes and Forests (5.2.1) of the upper midwest exhibit similar
patterns with respect to in the role of level and percentile in identifying regions not likely to meet
alternative standards, although there are considerably fewer cases compared to the regions in the
Coastal Plains and Appalachians.
       In the mountainous west, the Sierra Nevada (6.2.12), Idaho Batholith (6.2.15) and the
Cascades  (6.2.7) ecoregions likely do not meet alternative standards in fewer cases relative to
eastern regions, with the Sierra Nevada ecoregion exhibiting relatively greater sensitivity
compared to all western regions. Only in the upper part of the ranges of level and percentile do
regions in the northern and central Rockies likely not meet alternative standards.
       In considering these findings, it is clear that the  standard as defined by the AAI behaves
in an intuitively logical manner.  That is, an increase in ecoregions likely not to meet the
standard is associated with higher alternative levels and percentiles, both of which contribute to a
lower regionally representative critical load.  Moreover, the areas of known adverse aquatic
acidification effects are identified, mostly in high elevation regions or in the northern latitudes —
the Adirondacks, Shenandoahs, northern midwest lakes and the mountainous west.  These results
reflect the first application of a nationwide model that integrates water quality and atmospheric
processes at a national scale and provides findings that  are consistent with our basic
understanding of the extent of aquatic acidification across the U.S.  What is particularly
noteworthy is that this model is not initialized with a starting ANC based on water quality data,
which likely would result in a reproduction of water quality observations.  Rather, this standard
reflects the potential of the changes in atmospheric concentrations of NOy and SOx to induce
long-term sustained changes in surface water systems.  The fact that the patterns of adversity
based on applying this standard are commensurate with what is observed in surface water
systems provides confidence in the basic underlying formulation of the standard.
       However, the Coastal Plains and Appalachian mountain regions merit further inspection
as they stand out as areas with the largest relative exceedances from a national perspective. We

                                           7-61

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considered water quality data from these regions as well as an emissions sensitivity CMAQ
simulation to diagnose the behavior of these regions. The maps and tables in appendix D include
paired comparisons of the CMAQ 2005 and emissions sensitivity simulations.  The emissions
sensitivity simulation reflects domain-wide reductions in NOy and SOx emissions of 48% and
42%, respectively, relative to 2005 base year emissions.  We assume that this emissions
sensitivity simulation is indicative of future conditions.
       The emissions sensitivity results project that most the four Coastal Plains regions likely
not meeting alternative standards in the 2005 base year would likely continue not to meet the
standards in the future. In contrast, many of the regions that likely do not meet the alternative
standards based on recent air quality, especially at alternative levels of 20 and 35 |ieq/L,  would
likely meet such standards in the future year scenario for the Appalachian mountain regions.  It is
apparent that the AAI calculations are especially sensitive to changes in SOx emissions as the
Appalachian regions have the highest SOx concentrations and deposition rates (as discussed in
chapter 2), and we observe that the AAI equation responds as expected to  reductions in SOx.
On the other hand, the Coastal  Plains regions, especially the Atlantic Coastal Pine Barrens, have
extremely low critical loads relative to the eastern U.S., such that even the  least protective
alternative standards are likely not met despite significant projected reductions in SOx and NOy
emissions. The emissions sensitivity scenario is a prospective application of the standard, in the
sense that rules derived from the air quality management process result in reductions of NOy and
SOx emissions.  Expected emission changes over the next two decades should be far greater
than the 42 and 48%  SOx and NOy reductions used in this analysis, with a consequent further
reduction in ecoregions that would likely not meet alternative standards.
       Relative to other ecoregions,  water quality data for these Coastal Plains regions indicates
low natural base cation supply, low runoff rates and a large percentage of water bodies with
dissolved organic carbon (DOC) concentrations exceeding 5mg/L.  Because of both low  natural
base cation supply and runoff rates, indicating poor drainage, critical loads are near the bottom of
a national distribution (Table D-3). Elevated DOC and low base cation supply are indicative of
naturally acidic conditions where acidity is dominated by natural sources of organic acids and,
consequently, reductions in strong anions (MV and SO^2) resulting from reduced emissions
may provide only marginal benefits. Low base cation supply is not the cause of acidity, although
it is directly related to low ANC. In contrast, the Appalachian mountain regions generally have

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low DOC levels, average runoff rates, moderately low base cation supply and highly elevated
sulfate concentrations. Collectively, those attributes do not suggest naturally acidic conditions as
the availability of anthropogenic contributions of mineral acids is likely responsible for observed
low ANC values in those regions. Therefore, regions with limited base cation supply have
extremely low critical loads and when accompanied by elevated DOC levels are indicative of
naturally acidic systems.
       The Sierra Nevada region is an interesting case study as it has the lowest critical load
values nationally (Table D-2). Water quality data indicate extremely low sulfate, as expected
given the relatively low SO2 emissions in the western U.S.  Extremely low base cation supply
and low Neco, which mitigates the effect of nitrogen deposition, explain the low critical load
values.  Low Neco values appear to associate well with high elevation western U.S. regions,
perhaps reflecting the more arid and reduced vegetation density relative to eastern U.S. regions.
The proximity to high level nitrogen emissions combined with very low base cation supply
explains the cases where the Sierra region likely does not meet alternative standards.  Because
Neco values are low in the Sierras, the system responds effectively to reductions of NOx
emissions as illustrated in the maps  and tables of Appendix D. Although Neco affords protection
from the acidifying effects of nitrogen deposition, the availability of excessive nitrogen
neutralization capacity also means that reductions in nitrogen are not as effective as reductions in
SOx in reducing the  calculated AAI.
       In reviewing  these results, it is clear that the alternative combinations of level  and form
presented provide context for considering the impact of different standards.   Since the AAI
equation has been newly developed in this assessment, these exceedance examples help to
address the question  of whether the  AAI equation responds in a reasonable manner with regard
to identifying areas of concern and to prospective changes in atmospheric conditions likely to
result from future emissions reduction strategies.  In staffs view, the behavior of the AAI
calculations is both reasonable and explainable, which serves to increase our confidence in
considering a standard defined in terms of the AAI in this review.
       Further, these analyses provide additional insight in regard to categorizing the sensitivity
of ecoregions as acid sensitive or relatively non-acid sensitive, as introduced earlier in section
7.2.5.  In the earlier discussion, the Coastal Plains regions were highlighted as appearing
markedly  different than other acid sensitive regions located in mountainous and Northern latitude

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areas.  In presenting considerations that inform delineation of acid sensitive categories, we
interpret these exceedance analyses as suggesting that the Coastal Plains regions behave in a
markedly different manner relative to other acid sensitive regions.  In considering this
information, we note that the underlying assumption in developing an aquatic acidification
standard is that it is designed to afford protection from deposition-related risk beyond that which
arises from natural conditions. The lack of response of these regions to significant changes in
acidifying deposition would be consistent with the regions being categorized as relatively non-
acid sensitive.  The basis of this standard relies on the association between changes in emissions,
air concentrations, deposition and water quality. The concept of delineating relatively non-acid
sensitive areas is intended for those areas where this basic tenet of association is not adhered to.
Such is the case for extremely well buffered systems with high ANC, which was used in the
initial sensitivity categorizations discussed above in section 7.2, as well as for extremely poorly
buffered systems with low natural  base cation supply and naturally acidic conditions. This
analysis suggests that the following should appropriately be considered in delineating an
ecoregion as relatively non-acid sensitive: (1) the  level of natural base cation supply, (2) DOC
concentrations, (3) representative critical loads, and (4) responsiveness to deposition change.
We conclude that the combination of these considerations is more insightful than using just a
single attribute, as "natural" acidity is dependent on more than one variable.
       Consideration also should be given to other attributes, such as the dominant land use
(e.g., agriculture, commercial and residential development) and percentage of wild or protected
lands, both of which reflect the intent that this standard be focused on relatively pristine
environments.  In light of all  the above considerations, we conclude that it would be reasonable
and appropriate to categorize the Coastal plains ecoregions as relatively non acid-sensitive for
purposes of this standard.
       A categorization of relatively less-acid sensitive should not be interpreted as implying
that such areas would likely not receive benefits from programs designed to address acid
sensitive areas.  Since the relatively non-acid sensitive areas generally are in lower elevation
locations, they would generally benefit from the reduction of emissions designed to improve
nearby sensitive areas in two ways. First, because of the rural location of many acid  sensitive
areas, they often are located are in  so-called transport or regional air pollution corridors. The
regional behavior of NOy and SOx in the atmosphere is influenced strongly by emission

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strategies designed to reduce ambient ozone and PM2.5, both of which are influenced by the same
transport and atmospheric chemistry processes impacting acid sensitive regions. The emissions
sensitivity simulation illustrates the expected benefits associated with addressing regional scale
air pollution in a multiple pollutant context, recognizing that the CMAQ simulation originally
was used for assessing ozone and PM2.5. From an analogous hydrological perspective, the
reduction of acid anions flowing from higher elevation acid sensitive areas eventually translates
to reduced strong anion contributions, and therefore higher ANC, into the transitional plateaus,
plains and coastal areas.

7.6    SUMMARY OF SYSTEM UNCERTAINTIES
7.6.1   Overview
       This section summarizes discussions of results of analyses and assessments, presented
more fully in Appendices F and G, intended to address the relative confidence associated with
many of the individual and combined components of the linked atmospheric-ecological effects
system described throughout this chapter. These components include ecosystem effects; dose-
response relationships; underlying ecosystem sensitivity to acid deposition, biogeochemical,
atmospheric and deposition processes; and characterization of ecosystem services.
       Uncertainty and sensitivity analyses are used to inform the relative confidence in the
components and models that are used in defining the standard. Assessments of variability in the
data used to determine parameters of the standard increases the level of understanding about the
likelihood that alternative parameterizations of the standard will  achieve targeted levels of
protection when applied to sensitive ecosystems across the U.S.  Assessments of the sensitivity
of the calculated AAI to the components in the AAI equation can help demonstrate how
important uncertainty and variability in those components are in assessing the protection of
ecosystems provided by the standard. To evaluate the potential interactions between uncertain
and/or variable AAI components, a multifactor sensitivity analysis is also conducted. The ranges
of component values evaluated in the multifactor sensitivity assessment are guided by individual
variability and uncertainty analyses of specific components.  An additional objective of these
"confidence" related analyses and discussions is to help guide research and data collection
efforts intended to reduce uncertainty for future NAAQS reviews and implementation efforts.
Spatial and temporal variability analyses of AAI components are especially useful to  inform
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monitoring network design, spatial boundaries of acid sensitive areas, and consideration of multi-
year averaging periods.
       Significant emphasis is placed on evaluations of CMAQ due to the unique role that
atmospheric models hold in specifying terms in the AAI equation.  The AAI as developed in this
PA relies on CMAQ model simulations for both the initial characterization of reduced nitrogen
deposition and the deposition transformation ratios (TNox and TSox) which characterize the
relationships between atmospheric concentrations of NOy and SOx and deposition of nitrogen
and sulfur.   Included are interpretations of model evaluation results from the REA (EPA, 2009)
as well as more recent results related to wet deposition and the treatment of ammonia deposition.
Comparison of model results to observations provides a general sense of the confidence we have
that the models capture the spatial, temporal and compositional texture of the relevant
atmospheric and deposition species that drive the linked atmospheric-ecosystem processes. Both
model evaluation results and assessments of spatial and temporal variability can guide strategies
for  monitoring network design.  Sensitivity of CMAQ-derived deposition transformation ratios to
changes in atmospheric concentrations and variability over time provide insight into the stability
of these parameters that are used in a relatively static manner in the AAI, and into how well
alternative averaging times capture the overall spatial and temporal trends in the parameters.
       We evaluate the sensitivity of critical load modeling components by comparing dynamic
(MAGIC) and hybrid steady state  model results, looking at terminal results of MAGIC.   This
approach was viewed as a test of the more reduced form approximations used in steady state
modeling relative to more sophisticated treatment in MAGIC.
       For the purposes of this discussion, we characterize uncertainty regarding models and
their outputs as referring to the lack of knowledge regarding both the actual values of model
input variables (parameter uncertainty) and the model characterization of physical systems or
relationships (model uncertainty).  In any application, uncertainty is, ideally, reduced to the
maximum extent possible, but significant uncertainty often remains. It can be reduced by
improved measurement and improved model formulation. Model evaluation results provide
some insight into the relative uncertainty associated with  the ability of models to capture key
environmental state characteristics. Confidence regarding the fundamental science  supporting
causal determinations about the effects of acid deposition, and the translation of those effects
into ecosystem services and values is less amenable to quantification.

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       Sensitivity refers to the influence on modeled results due to perturbations in input
variables or change of process formulations.  Sensitivity analysis can provide a sense of how
important different parameters and inputs might be to the outcomes of interest, ie., the calculated
AAI value, but cannot by themselves indicate how important specific parameters actually are,
because they do not incorporate information on the range of parameter values or the likelihood
associated with any specific parameter value.  Sensitivity results in this assessment are intended
to provide insight into the relative stability of the AAI and confidence in modeled
parameterizations. Sensitivity analyses are especially useful in the absence of observed data to
challenge models.  For example, the NOy and SOx transference ratios are a model construct that
is difficult, if not impossible, to compare to observations.  The sensitivity of these ratios to
changing meteorology and emissions is evaluated in reference to the stability of these ratios
under changing conditions.  Low sensitivity here implies that the choice to use long-term
averages of modeled ratios is justified.  Sensitivity analyses also are used to discern the relative
influence  on calculated AAI values of other parameters in the AAI equation. Toward that end,
elasticity analyses were applied to determine the relative sensitivity of AAI results associated
with individual and combined AAI parameters.  A Monte Carlo type simulation was also
conducted to inform characterization of overall uncertainty associated with the AAI equation.
       Variability refers to the heterogeneity in a population or variable of interest that is
inherent and cannot be reduced through further data collection and research. In the context of
the AAI, characterization of variability  can be used to guide the design of an appropriate
monitoring network.
7.6.2  Summary of results and conclusions
       Uncertainty and natural variability exist in all of the components of the AAI developed in
this PA, and should be considered in establishing a standard for aquatic acidification.  A
summary  of the relative uncertainties of these components is provided in Table 7-3. On balance,
the confidence level in the information  and processes associated with the linkages from
ecological effects to atmospheric conditions through deposition and ecosystem modeling is very
high.
       An analysis of the cumulative effects of uncertainty on the AAI was conducted and is
described in Appendix G.  In summary, this included bootstrapping analyses of the parameters in
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the AAI equation to translate error in individual measurements to the regional values used in the
equation.  The parameters that are averages of grid-level CMAQ modeled values, Ndep and NHx,
had bootstrapped uncertainty values of approximately +20% (Figures G-l and G-5).  The
transference ratios include two different grid-level CMAQ modeled values and had much higher
uncertainty, exceeding 100% (Figures G-3 and G-4).  The calculation ofNeco also includes two
different input values, the CMAQ derived Ndep values and lake-specific nitrogen leaching
values.  The Neco results also had high uncertainty values, ranging from -65% to approximately
200% (Figure G-2). The critical load value for the region is affected by the Q and BCo values at
the individual lakes within a region. Uncertainty in these parameters gave a regional uncertainty
range for the critical load of+35% (Figure G-6).
       The results of the bootstrapping analyses were used to complete a cumulative analysis of
uncertainty in a subsequent Monte Carlo style analysis. This analysis is illustrated in the form of
the tradeoff curve for the concentrations of NOy and SOx (Figure  G-7).  The results in the two
regions analyzed were similar. There was a range of uncertainty, with 50% of the distribution
within +20% of the observed value.  Most importantly, the mean value of the results was very
close to the observed value  in both regions. This indicates that there is no systematic bias in the
results despite what can be relatively high levels of uncertainty in the input parameters.
       The considerable body of evidence is conclusive with regard to causality between aquatic
acidification and biological and ecological effects. Confidence in  the linkage associating aquatic
acidification and ANC is extremely high, as the aquatic chemistry describing this relationship,
while nonlinear, is relatively simple with regard to chemical species and reactions. The
relationships between deposition and ANC, while complicated by  a variety of biogeochemical
and hydrological processes  and data requirements within watersheds, are well established and the
critical load models have been thoroughly vetted through the scientific community with a
demonstrated level of successful evaluation. The linkages between ambient concentrations of
relevant species and deposition are best handled through air quality modeling systems like
CMAQ. The relationship between concentrations and deposition is well characterized by these
models, which are constrained by mass balance principles. While much of the physical and
chemical processing that determines concentrations and consequent deposition is interwoven
with numerous fundamental processes characterizing mass transport and atmospheric chemical
oxidation, the science is  relatively mature with years of applications and continued evolution of

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the models.  The specific processes guiding nitrogen and sulfur chemistry and deposition are
relatively simple. More challenging is the ability to parameterize processes at the air-surface
interface which guide the estimation of deposition velocities and the re-emission of certain
species, as well as many of the area-wide natural processes and agricultural practices which
influence emissions of oxidized and reduced forms of nitrogen.
       The variety of uncertainty, variability, and sensitivity analyses included in Appendices F
and G have been conducted under the assumption that the basic model construct is well
established, as discussed immediately above.  Throughout these discussions there is no apparent
directional bias in the uncertainty regarding the biological, chemical and physical processes
incorporated in the AAI.  From the perspective of valuation of ecosystem services, the estimates
generally are believed to be biased low, meaning the values of reaching a target level of
protection are underestimated.  However, quantification of these values is perhaps the most
uncertain of all aspects considered.
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Table 7-3. Summary of Qualitative Uncertainty Analysis of Key Components of the AAI
Source
Description
Potential influence of
uncertainty in element
Direction
(negative
implies less
relative
protection)
Magnitude
Knowledge-
Base
uncertainty
Comments
Major elements (and sub-models) of the ecological effects to ambient concentration framework
Biological/ecosystem
response to
acidification
Linkage between
direct acidification
species and
ecological indicator
(ANC)
Linkage between
ecological indictor
and adverse
ecological effects
Deposition to ANC
linkage through
Critical Load
approach
Clear associations between
aquatic acidification (pH,
elevated Al) and adverse
ecosystem effects (fish
mortality, decreased species
diversity)
The relationships across ANC,
pH and dissolved Al are
controlled by well defined
aquatic equilibrium chemistry
Direct nonlinear associations
between ANC and fish
mortality and species diversity
Mass-balance Steady State
critical load model is applied to
determine critical load values.
MAGIC model is used to
validate steady State model.
The Steady State critical load
model formulation is used as
the foundation for deriving the
AAPI equation.
Both
Both
Both
Both
Low
Low
Low-
medium
Low
Low
(regionally)
Low
Low
Low
The ecosystem level responses are well studied at regional
levels. The uncertainty increases at larger scales due to an
increasing number of factors influencing the patterns (e.g.
latitudinal species gradient, specie-area relationships).
ANC is the preferred ecosystem indicator as it has a direct
relationship with pH and the deposition species relevant to the
NOx/SOx standard.
Although the pH dependency on ANC is nonlinear, it is always
directionally consistent. In extremely low and high ANC
environments the relationship is of minimal value as
catchments are in relatively "less sensitive" regimes due to
natural conditions or extreme anthropogenic influence (i.e.,
acid mine drainage). In sensitive areas of concern the
relationship essentially is similar to the relationships between
direct acidification species and adverse effects.
The model formulation is well conceived and based on a
substantial amount of research and applications available in the
peer reviewed literature. There is greater uncertainty
associated with the availability of data to support certain model
components.
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       Source
         Description
                                                        Potential influence of
                                                       uncertainty in element
 Direction
 (negative
implies less
  relative
protection)
Magnitude
Knowledge-
    Base
uncertainty
                        Comments
Atmospheric
concentrations to
deposition
Deposition is a direct function
of ambient concentration,
influenced by several
processes, and handled in the
AAPI through air quality
modeling.
   Both
 Low
    Low
The model design is appropriate given the spatial and temporal
complexities that influence deposition velocity, as well as the
variety of atmospheric species that generally are not measured.
Greater uncertainty resides in the information (e,g., ammonia
emissions) driving these calculations and availability of
observations to evaluate model behavior.
Ecological indicator
to changes in the
value of ecosystem
services
Definitions of public welfare
may include economic
considerations, based on the
tradeoffs people would make to
avoid the negative impacts of
acidification, through effects on
the values of ecosystem
services. Empirical estimates
of valuation for limited
ecosystem service categories
are used to inform the
discussions of adversity
associated with alternative
ANC  levels.
 Negative
 Medium-
 high
Low-medium
There are many studies that estimate the value of increasing
services that may be affected by changes in acidification and
eutrophication.  However, few of these studies focus on the
particular impact of acidification and eutrophication on the
quality of these services and preferences for avoiding these
impacts.

Those studies that do are often limited to analyzing the impacts
on a narrow population or particular change in environmental
quality. The monetized benefits to fishers and to New York
residents for ecosystem improvements in the Adirondacks
associated with improvements to the ecological indicator are
significant underestimates of the total benefits in the U.S. This
is because those living outside New York would value
improvements to the Adirondacks and similar natural
environments elsewhere.

The methodologies used in the studies that underlie the
estimates of the value of changes in ecosystem services in the
Adirondacks region are sound and have been subject to peer
review. The method of aligning the improvements valued in
the Banzhaf et al. study with estimates of eliminating current
damages leads to may lead to an over or underestimate of the
benefits.  The range of this difference is difficult to know a
priori, but the total improvements in the share of lakes that
improve above an ANC threshold of 20 ueq/L are consistent.
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      Source
         Description
                                                       Potential influence of
                                                      uncertainty in element
 Direction
 (negative
implies less
  relative
protection)
Magnitude
Knowledge-
   Base
uncertainty
                       Comments
                                                Sub-components and data of individual models
                                                                Atmospheric Components
Deps
Annual deposition of sulfur
mass from dry deposition of
(SO2 and SO4) and wet SO4
derived from CMAQ 12km
horizontal grid resolution
averaged over 5 years	
   Both
   Low
    Low
The treatment of SOx deposition in EPA air quality models has
evolved over the last two decades. There is general consensus
that the overall mass balance of S is treated well with
difficulties in spatial pairing of observations and modeled
results of wet deposition. This spatial pairing has improved
with the more recent PRISM adjustments.	
                      Annual deposition of oxidized
                      nitrogen mass from dry
                      deposition of (all NOy species)
                      and wet NO3 derived from
                      CMAQ 12 km horizontal grid
                      resolution averaged over 5
                      years
                                  Both
                 Low
              Low-medium
               The treatment of oxidized nitrogen deposition in EPA air
               quality models has evolved over the last two decades.  There
               is general consensus that the overall mass balance of oxidized
               N is treated well. However, the broad range of deposition
               velocities across NOy species, and especially uncertainties
               regarding the deposition of significant species such as NO2
               pose ongoing challenges. Similarly, a shortage of NOy species
               measurements as well a lack of techniques to directly measure
               dry deposition impede progress  on improving parameterization
               of N dry deposition.	
                      Annual deposition of reduced
                      nitrogen mass from dry
                      deposition of (NH3 and SO4)
                      and wet NH4 derived from
                      CMAQ 12km horizontal grid
                      resolution averaged over 5
                      years
                                  Both
                 Low
                Medium
               NHx deposition also is quantified through CMAQ applications.
               The well dispersed nature of agricultural based emissions that
               are influenced strongly by meteorological and surface /soil
               characteristics continues to challenge characterization of
               ammonia emissions.  Recent incorporation of a bi-directional
               flux process in CMAQ improves consistency with available
               scientific understanding and yields improved time and space
               pairing of limited observations with model results.  A lack of
               both ammonia and ammonium ambient observations continues
               to compromise our ability to characterize uncertainty in our
               treatment of NHx. As with all dry deposition estimates,
               technologies for direct measurements are not available
               routinely.  Both NHx deposition and NOx deposition are
               assigned low values of magnitude based on a general
               dominating role of sulfur deposition.	
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      Source
         Description
                                                        Potential influence of
                                                       uncertainty in element
 Direction
 (negative
implies less
  relative
protection)
Magnitude
Knowledge-
    Base
uncertainty
                       Comments
Wet deposition
(genetically - N and
S species)
Wet component of total
deposition as described in the
Dep terms, above
   Both
   Low
    Low
Wet deposition remains an attribute of relatively high
confidence based on the ability to directly measure chemical
components in precipitation samples. However, given the
stochastic nature of precipitation, models have a difficult time
in matching observations. The use of 5 year averages and
post-processing PRISM adjustments have reduced uncertainty
in spatial pairing of observations and modeled estimates.	
Dry deposition
(generically - N and
S species)
Dry component of total
deposition as described in the
Dep terms, above
   Both
 Medium
Medium-high
The absence of direct dry deposition measurements combined
with the significant variability in the parameters that influence
dry deposition velocity reduces the confidence level in dry
deposition relative to wet deposition.	
Deposition
Transference Ratios
CMAQ derived ratio of total
oxidized deposition to
concentration averaged over
one year
   Both
   Low
  Unknown
Transference ratios enable the connection between deposition
and the policy relevant ambient air indicators, NOy and (SO2 +
SO4). They are strictly a model construct and cannot be
evaluated in a traditional model to observation context. The
low sensitivity of these ratios to emission changes and inter
annual meteorology combined with low spatial variability
indicate that these ratios are necessarily stable.	
                      Ambient concentrations of
                      NOy through observations.
                                 Negative
                  Low
              Low-medium
                Adequate spatial coverage of NOy observations does not exist,
                but will be addressed in the proposed rule.  The monitoring
                technology only over the last 5 years has been perceived as
                "routine" based on incorporation in the NCore network.
                However , FRM status for NOy instruments currently is not
                available.  The negative bias direction is a standard caveat to
                any instrument relying on internal air stream conversion of
                atmospheric species prior to detection.
                      Ambient concentrations of
                      NOy through observations.
                                   Both
                  Low
                  Low
                A lack of adequate spatial coverage is the primary concern for
                SO2 + SO4 observations. FRM status is not available for SO4;
                although the long track record of accurate and precise
                CASTNET FP measurements indicates that achieving FRM
                status is a low hurdle.
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      Source
         Description
                                                        Potential influence of
                                                       uncertainty in element
 Direction
 (negative
implies less
  relative
protection)
Magnitude
Knowledge-
   Base
uncertainty
                        Comments
                                                                  Ecosystem Components
BCn
Pre-industrial base cation
concentrations
 Negative
 Medium-
   high
   High
Both the F-factor approach and process based MAGIC
modeling were used to generate BC0* Excellent agreement
between both approaches was established in the Shenandoah
streams. The more comprehensive data requirements of
MAGIC limit its widespread use to the Adirondacks, although
for consistency the F-factor approach was applied nationwide.
The analyses also illustrated greater divergence at higher
critical loads, or areas with greater acid buffering capacity and
high base cation levels. These conditions often are screened
out of our population distribution analyses,  and when included
do not affect the location within the distribution of the more
sensitive water bodies.  Since MAGIC (the  preferred approach)
tends to overestimate BC0*relative to the F factor approach, and
the F-factor is more widely applied nationally, the BC0*
estimates are viewed as conservative leading to a slight
positive bias in estimating critical loads. Although we have
many modeled estimates of BC0*, there is a lack of direct
measurements of BC weathering rates.	
Neco
                                  Positive
                  Low
                 Medium
               The term Neco, as defined, has a relatively medium confidence
               level and is a direct function of the uncertainty inherent in the
               deposition estimates from CMAQ and surface measurements
               of NO3. However, this "measurement" difference approach
               reflects the average of all influencing processes (dinitrification,
               uptake, and immobilization) over the time period of
               measurements.  Consequently, there is an inherent assumption
               of a relatively static system (Neco is applied in a steady state
               model) that generally is not tested.  In concept, a true steady
               state vision of Neco would be based on a mature forested
               ecosystem. The relative bias of Neco is related, largely, to the
               relative productivity of the forest.  The challenge  in
               determining any potential bias in Neco is to determine the
               relative "maturation age" of an ecosystem which requires	
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      Source
         Description
                                                        Potential influence of
                                                       uncertainty in element
 Direction
 (negative
implies less
  relative
protection)
Magnitude
Knowledge-
   Base
uncertainty
                       Comments
                                                                                                  knowledge of future land use activities. In areas of high land
                                                                                                  use restrictions of a recovering forest, Neco would be assumed
                                                                                                  to be overestimated.  The relative magnitude of Neco often is
                                                                                                  mitigated by the dominance of SOx in controlling acidification
                                                                                                  processes in many systems.  Furthermore, it is unclear to what
                                                                                                  extent any stored N will be released back into the system,
                                                                                                  which is assumed to not occur in the linked system model.
Q
Annual runoff rate
(distance/time) for a catchment.
   Both
   Low
   High
Data used to calculate Q was compiled in 1985.  Streamflow
data were collected  at over  12,000 gauging stations during
1951-80; 5,951 stations were  selected for the analysis.  See
Gebert et al. (1987) for a complete description  of how the
runoff was determined from the streamflow data.  Appropriate
maps of the data can show the geographical distribution of
runoff in tributary  streams for  the years  1951-80 and can
describe  the magnitudes and variations of runoff nationwide.
The data was prepared to reflect the runoff of tributary streams
rather  than   in  major  rivers in  order  to  represent  more
accurately the local  or small  scale variation in  runoff with
precipitation and other geographical characteristics.

Gerbert, W.A., Graczyk, D.J., and Krug, W.R., 1987, Average
annual runoff in the United States, 1951-80:  U.S. Geological
Survey   Hydrologic   Investigations  Atlas   HA-710,   scale
1:7,500,000.	
DOC
Surface water dissolved
organic carbon
 Negative
   Low
  Medium
Water bodies with high DOC levels (> 10mg/l) were screened
out of the critical load calculations in order to avoid naturally
acidic systems.  However, the inherent assumption of  ANC =
Xstrong CA - £strong AN does not explicitly account for
contributions of weak organic acids.  Consequently, a  small
positive bias pervades the critical load calculations (i.e., the CL
estimates are high).   The knowledge base value of M reflects
a general shortage of DOC data.	
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7.7    SUMMARY OF STAFF CONCLUSIONS ON SECONDARY STANDARDS FOR
       OXIDES OF NITROGEN AND SULFUR
       This section summarizes staff conclusions with regard to the adequacy of the current NO2
and SO2 secondary standards and potential alternative standards that are appropriate to consider
to provide requisite protection from adverse public welfare effects, including effects on sensitive
ecosystems, associated with the presence and deposition of oxides of nitrogen and sulfur in the
ambient air. In reaching these conclusions, staff has considered these standards in terms of the
basic elements of the NAAQS: indicator, averaging time, form, and level (as discussed above in
sections 7.1 to 7.4, respectively). Staff conclusions are based on the available scientific and
technical information as assessed and presented in the ISA (US EPA, 2008), the REA (US EPA,
2009), and as summarized and interpreted throughout this documents and its appendices. In so
doing, we have considered the advice of CAS AC and public comments on earlier drafts of this
document.
       In this assessment, we emphasize a policy approach that incorporates a multi-pollutant,
multi-media framework, taking into consideration the combined effects of oxides of nitrogen and
oxides of sulfur and the linkages between relevant atmospheric processes and associated
ecosystem effects.   As such, we have taken into account both evidence-based and impact
assessment-based considerations to inform our conclusions related to the adequacy of the current
NO2 and SO2 secondary standards and alternative standards that are appropriate for consideration
in this review.  In so doing, we are seeking to identify as broad an array of policy options as is
supportable by the available information, recognizing that the selection of a specific approach to
reaching final decisions on secondary standards for oxides of nitrogen and sulfur will reflect the
judgments of the Administrator as to standards that are requisite to protect the public welfare
from adverse effects associated with the  presence of oxides of nitrogen and sulfur in the ambient
air.
       We recognize that selecting from among alternative standards will necessarily reflect
consideration of the qualitative and quantitative uncertainties inherent in the relevant evidence
and in the quantitative impact assessment of exposure and risks to sensitive ecosystems. In
reaching staff conclusions on alternative standards that are appropriate to consider, we are
mindful that the CAA requires secondary standards to be set that are requisite to protect public
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welfare from known and anticipated adverse effects, such that the standards are to be neither

more nor less stringent than necessary.

       Based on the currently available information, staff reaches the following conclusions

regarding secondary standards for protecting against adverse public welfare effects, including

effects on sensitive ecosystems, associated with the presence and deposition of oxides of
nitrogen and sulfur in the ambient air:

(1) With regard to the adequacy of the current standards, currently available scientific evidence
   and assessments clearly call into question the adequacy of the protection afforded by the
   current NC>2 and SC>2 secondary standards from deposition-related effects on sensitive
   ecosystems related to oxides of nitrogen and sulfur in the ambient air, including acidification
   and nutrient enrichment in aquatic and terrestrial ecosystems.  In  addition, the elements of the
   current NC>2 and SC>2 standards are not ecologically relevant, and are thus not appropriate, for
   standards that are designed to provide such protection. Nonetheless, based on the current
   evidence, the current standards likely do afford adequate protection from the direct effects
   involving injury to vegetation associated with atmospheric exposure to oxides of nitrogen and
   sulfur.
       (a) Thus, consideration should be given to  establishing a new ecologically relevant
           standard(s) to provide increased protection from deposition-related  effects of oxides
           of nitrogen and  sulfur  on sensitive ecosystems.
       (b) Consideration should also be given to retaining the current NC>2 and SC>2 secondary
           standards to continue to provide protection from the direct effects of oxides of
           nitrogen and sulfur on vegetation.
(2) With regard to establishing a  new ecologically relevant standard(s), consideration should be
   given to establishing  a multi-pollutant standard that addresses the combined effects of oxides
   of nitrogen and oxides of sulfur specifically to  provide increased  public welfare protection
   from aquatic acidification in sensitive ecosystems. This conclusion is based in general on the
   evaluation and assessments in the ISA and REA showing that both oxides of nitrogen and
   sulfur are major contributors to aquatic acidification and that acidification of aquatic
   ecosystems is best characterized and understood in terms of the combined rather than
   individual effects of oxides of nitrogen and sulfur. In addition, there is a well  developed
   body of scientific evidence linking the deposition of ambient oxides of nitrogen and sulfur to
   acidification in sensitive aquatic ecosystems and showing that a significant number of water
   bodies currently experience levels of acidification associated with the deposition of
   atmospheric nitrogen and sulfur that could reasonably be judged to be adverse from a public
   welfare perspective in areas where the current standards were met.
       (a) While there are  other important ecosystem effects attributable to deposition of oxides
           of nitrogen and/or oxides of sulfur, such as terrestrial acidification and nutrient
           enrichment of aquatic  and terrestrial systems, we conclude that the available
           information and assessments are only sufficient to support the development of a
           national standard specifically to address aquatic acidification at this time.
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       (b) Nonetheless, we recognize an aquatic acidification standard as developed in this
          assessment is likely to provide some degree of co-protection for these other
          deposition-related effects, particularly for terrestrial acidification, in at least some
          acid sensitive watersheds.
(3) With regard to ambient air indicators for an aquatic acidification standard that would address
   the combined contributions of oxides of nitrogen and sulfur, consideration should be given to
   using total reactive oxidized nitrogen, NOy, as the indicator for oxides of nitrogen and the
   sum of gaseous sulfur dioxide (862) and paniculate sulfate (SO/t), referred to in this
   assessment as SOX, as the indicator for oxides of sulfur. This conclusion takes into
   consideration the available evidence that demonstrates a strong linkage between
   concentrations of NOy and SOX in the ambient air, the deposition of nitrogen and sulfur from
   NOy and SOX, and acidification effects in aquatic ecosystems.
       (a) Consideration could also be given to defining the indicator for oxides of nitrogen as
          some subset of NOy species, including those NOy species that deposit relatively more
          quickly than  other species and/or those that comprise the dominant mass of NOy.
          Based on our assessment, however, we conclude that the advantages of using total
          aggregated NOy make it the preferred choice.
       (b) In considering an indicator for oxides of nitrogen, we recognize that aquatic
          acidification  results from and is best understood in terms of the deposition of total
          nitrogen, in both oxidized and reduced forms. Since the pollutant that is the focus of
          this review is oxides of nitrogen, not reduced forms of nitrogen, we conclude that it is
          appropriate to consider reduced forms of nitrogen separately, as a factor in the form
          of the standard, rather than as part of the indicator of the standard.
(4) With regard to the form of such a multi-pollutant, deposition-related standard, consideration
   should be given to an ecologically relevant form that  characterizes the relationships between
   the ambient air indicators for oxides of nitrogen and sulfur, the related deposition of nitrogen
   and sulfur, and the associated aquatic acidification effects in terms of a relevant ecological
   indicator.  Based on  the available information and assessments, consideration should be
   given to using acid neutralizing capacity (ANC) as the most appropriate ecological indicator
   for this purpose, in that it provides the most stable metric that is highly associated with the
   water quality properties that are directly responsible for the principal adverse effects
   associated with aquatic acidification: fish mortality and reduced aquatic species diversity.
   We have developed such a form, termed an aquatic acidification index (AAI), using a simple
   equation to calculate an AAI value in terms of the ambient air indicators of oxides and
   nitrogen and sulfur and the relevant ecological and atmospheric factors that modify the
   relationships between the ambient air indicators and ANC.  Recognizing the spatial
   variability of such factors across the U.S., we conclude it is appropriate to divide the country
   into ecologically relevant regions, characterized as acid-sensitive or relatively non-acid-
   sensitive, and specify the value of each of the factors in the AAI equation for each such
   region.
   With regard to approaches to defining such ecologically relevant regions, consideration
   should be given to using Omernik ecoregions, level III, as the appropriate set of regions over
   which to define the AAI.  There are 84 such ecoregions that cover the continental U.S. This
   set of ecoregions is based on grouping a variety of vegetation, geological, and hydrological
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attributes that are directly relevant to aquatic acidification assessments and that allow for a
practical application of an aquatic acidification standard on a national scale.
With regard to an equation that would define the AAI, consideration should be given to the
following equation:

           AAI = Fl - F2 - F3[NOy] - F4[SOX]

Factors Fl through F4 would be defined for each ecoregion by specifying ecoregion-specific
values for each factor based on monitored or modeled data that are representative of each
ecoregion. The Fl factor is also defined by a target ANC value. More specifically:
    (a) Fl reflects a relative measure of an ecosystem's ability to neutralize acidifying
       deposition. The value of Fl for each ecoregion would be based on a representative
       critical load for the ecoregion associated with a single national target ANC level, as
       well as on a representative runoff rate.  The representative runoff rate, which is also
       used in specifying values for the other factors, would be the median value of the
       distributions of runoff rates within the ecoregion. The representative critical load
       would be derived from a distribution of critical loads calculated for each water body
       in the ecoregion for which sufficient water quality and hydrology data are available.
       The representative critical load would be defined by selecting a specific percentile of
       the distribution.
       In identifying a range of percentiles that are appropriate to consider for this purpose,
       we have considered regions categorized as acid sensitive separately from regions
       categorized as relatively non-acid sensitive. In delineating these categories,
       consideration should be given to alternative approaches that take into account a range
       of relevant ecological and atmospheric factors. For acid  sensitive regions, we
       conclude that consideration should be given to selecting a percentile value from
       within the range of the 70th to the 90th percentile.  The lower end of this range was
       selected to be appreciably above the median value so as to ensure that the critical load
       would be representative of the population of relatively more acid sensitive water
       bodies within the region, while the upper end was selected to avoid the use of a
       critical load from the extreme tail of the distribution which is subject to a high degree
       of variability and potential outliers.  For relatively non-acid  sensitive regions, we
       conclude that consideration should be given to selecting the  50th percentile to best
       represent the distribution of water bodies within such a region, or alternatively to
       using the median critical load of all relatively non-acid sensitive areas, recognizing
       that such areas are far less frequently evaluated than acid sensitive areas. Using
       either of these approaches would avoid characterizing a generally non-acid-sensitive
       region with a critical load that is representative of relatively  acid sensitive water
       bodies that may exist within a generally non-acid sensitive region.
    (b) F2 reflects the deposition of reduced nitrogen. Consideration should be given to
       specifying the value of F2 for each region based on the averaged modeled value
       across the region, using national CMAQ modeling that has been conducted by EPA.
       Consideration could also be given to alternative approaches  to specifying this value,
       such as allowance for the use of air quality modeling conducted by States using more
       refined model inputs.
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       (c) F3 and F4 reflect transference ratios that convert ambient air concentrations of NOy
          and SOX, respectively, into related deposition of nitrogen and sulfur.  Consideration
          should be given to specifying the values for F3 and F4 for each region based on
          CMAQ modeling results averaged across the region. We conclude that specifying the
          values or the transference rations based on CMAQ modeling results alone is preferred
          to an alternative approach that combines CMAQ model estimates with observational
          data.
       (d) The terms [NOy] and [SOX] reflect ambient air concentrations measured at monitoring
          sites within each region.
(5) With regard to averaging time, consideration should be given to averaging calculated annual
   AAI values over 3 to 5 years to provide reasonable stability in the resulting index value, in
   light of the relatively high degree of interannual variability expected in an index that is
   strongly related to the amount and pattern of precipitation that occurs within a region from
   year to year.
(6) With regard to the level of a standard based on the above indicators, alternative forms, and
   averaging times, consideration should be given to a level within the range of 20 to 75 |ieq/L.
   In reaching this conclusion, staff has considered the available information that links specific
   ANC levels to various types of acidification-related effects, and the uncertainties inherent in
   such linkages, and the severity of such effects, in sensitive ecosystems, as well as the extent
   to which such effects could reasonably be judged to be important from a public welfare
   perspective. This range also reflects consideration of the extent to which such a standard
   would protect against not only long-term but also episodic acidification, as well as the time
   lag in ecosystem response to changes in deposition that may result from such a standard.
   Relatively more protection from both long-term and episodic acidification would be provided
   by a standard in the mid- to upper part of this range, which would also accelerate the time
   frame in which the target ANC level would likely be reached in some sensitive ecosystems.
   This range also encompasses target ANC values that have been established by various States
   and regional and international organizations to protect against acidification of aquatic
   ecosystems.
   Based on the evidence and assessments in the ISA and REA, we conclude that a target ANC
   value of 20 jieq/L  is a reasonable lower end of this range, so as to protect against chronic
   acidification-related adverse impacts on fish populations which have been characterized as
   severe at ANC values below this level.  Further, we conclude that a target ANC value of 75
   |ieq/L is a reasonable upper end of this range in recognition that the potential for additional
   protection at higher ANC values is substantially more uncertain in light of evidence that
   acidification-related effects are far less sensitive to increases in ANC above this value.
(7)  An aquatic acidification standard, as defined above, would be interpreted as follows: the
   standard would be met at a monitoring site when the measured annual-average concentrations
   of NOy and SOX are such that the value of the  annual AAI, averaged over 3 to 5 years, is
   equal to or greater than the level of the standard, when using the region-specific values of
   factors Fl through F4 for the ecoregion in which the monitor is located.
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7.8 REFERENCES

Hemond. 1990. Acid Neutralizing Capacity, Alkalinity, and Acid-Base Status of Natural Waters
     Containing Organic Acids, Environ Sci. Technol, 24, 1486-1890.
Omernik, J.M. 1987. Ecoregions of the conterminous United States. Map (scale 1:7,500,0000).
     Annals of the Association of American Geographers 77:118-125.
Seinfeld JH; Pandis SN. (1998). Atmos Chem Phys: From Air Pollution to Climate Change. New
     York, NY: John Wiley-Interscience Publishers.
Smyth SC, W. Jiang and H. Roth, 2008, A comparative performance evaluation of the AURAMS
     and CMAQ air quality modelling systems. Atmos Envir 43:1059-1070.
Stumm, W. and J.J. Morgan. 1981. Aquatic Chemistry, Wiley-Interscience, 2nd edition.
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United States                                     Office of Air Quality Planning and Standards                     Publication No. EPA-452/R-1 l-005a
Environmental Protection                          Health and Environmental Impacts Division                                         February 2011
Agency                                                 Research Triangle Park, NC

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