enrichment and one in Little Rock Lake, Wisconsin for aquatic nutrient
enrichment.
2. Key Findings
    In summary, based on case study analyses, the REA concludes that
known or anticipated adverse ecological effects are occurring under
current conditions and further concludes that these adverse effects
continue into the future. Key findings from the air quality analyses,
acidification and nutrient enrichment case studies, as well as general
conclusions from evaluating additional welfare effects, are summarized
below.

a. Air Quality Analyses
    The air quality analyses in the REA encompass the current emissions
sources of nitrogen and sulfur, as well as atmospheric concentrations,
estimates of deposition of total nitrogen, policy-relevant background,
and non-atmospheric loadings of nitrogen and sulfur to ecosystems, both
nationwide and in the case study areas. Spatial fields of deposition
were created using wet deposition measurements from the NADP National
Trends Network and dry deposition predictions from the 2002 CMAQ model
simulation. Some key conclusions from this analysis are:
    (1) Total reactive nitrogen deposition and sulfur deposition are
much greater in the East compared to most areas of the West.
    (2) These regional differences in deposition correspond to the
regional differences in oxides of nitrogen and SO2
concentrations and emissions, which are also higher in the East. Oxides
of nitrogen emissions are much greater and generally more widespread
than NH3 emissions nationwide; high NH3 emissions
tend to be more local (e.g., eastern North Carolina) or sub-regional
(e.g., the upper Midwest and Plains states). The relative amounts of
oxidized versus reduced nitrogen deposition are consistent with the
relative amounts of oxides of nitrogen and NH3 emissions.
Oxidized nitrogen deposition exceeds reduced nitrogen deposition in
most of the case study areas; the major exception being the Neuse
River/Neuse River Estuary Case Study Area.
    (3) Reduced nitrogen deposition exceeds oxidized nitrogen
deposition in the vicinity of local sources of NH3.
    (4) There can be relatively large spatial variations in both total
reactive nitrogen deposition and sulfur deposition within a case study
area; this occurs particularly in those areas that contain or are near
a high emissions source of oxides of nitrogen, NH3 and/or
SO2.
    (5) The seasonal patterns in deposition differ between the case
study areas. For the case study areas in the East, the season with the
greatest amounts of total reactive nitrogen deposition correspond to
the season with the greatest amounts of sulfur deposition. Deposition
peaks in spring in the Adirondack, Hubbard Brook Experimental Forest,
and Kane Experimental Forest case study areas, and it peaks in summer
in the Potomac River/Potomac Estuary, Shenandoah, and Neuse River/Neuse
River Estuary case study areas. For the case study areas in the West,
there is less consistency in the seasons with greatest total reactive
nitrogen and sulfur deposition in a given area. In general, both
nitrogen and/or sulfur deposition peaks in spring or summer. The
exception to this is the Sierra Nevada Range portion of the MCF Case
Study Area, in which sulfur deposition is greatest in winter.
b. Deposition-Related Aquatic Acidification
    The role of aquatic acidification in two eastern United States
areas--northeastern New York's Adirondack area and the Shenandoah area
in Virginia--was analyzed in the REA to assess surface water trends in
SO42- and NO3-
concentrations and ANC levels and to affirm the understanding that
reductions in deposition could influence the risk of acidification.
Monitoring data from the EPA-administered TIME)/LTM programs and the
EMAP were assessed for the years 1990 to 2006, and past, present and
future water quality levels were estimated using both steady-state and
dynamic biogeochemical models.
    Although wet deposition rates for SO2 and oxides of
nitrogen in the Adirondack Case Study Area have reduced since the mid-
1990s, current concentrations are still well above pre-acidification
(1860) conditions. The MAGIC modeling predicts
NO3- and SO42- are 17- and
5-fold higher today, respectively. The estimated average ANC for 44
lakes in the Adirondack Case Study Area is 62.1 [mu]eq/L (15.7 [mu]eq/L); 78 percent of all monitored lakes in the
Adirondack Case Study Area have a current risk of Elevated, Severe, or
Acute. Of the 78 percent, 31 percent experience episodic acidification,
and 18 percent are chronically acidic today.
    (1) Based on the steady-state critical load model for the year
2002, 18 percent, 28 percent, 44 percent, and 58 percent of 169 modeled
lakes received combined total sulfur and nitrogen deposition that
exceeded critical loads corresponding to ANC limits of 0, 20, 50, and
100 [mu]eq/L respectively.
    (2) Based on a deposition scenario that maintains current emission
levels to 2020 and 2050, the simulation forecast indicates no
improvement in water quality in the Adirondack Case Study Area. The
percentage of lakes within the Elevated to Acute Concern classes
remains the same in 2020 and 2050.
    (3) Since the mid-1990s, streams in the Shenandoah Case Study Area
have shown slight declines in NO3 and
SO42- concentrations in surface waters. The ANC
levels increased from about 50 [mu]eq/L in the early 1990s to >75
[mu]eq/L until 2002, when ANC levels declined back to 1991-1992 levels.
Current concentrations are still above pre-acidification (1860)
conditions. The MAGIC modeling predicts surface water concentrations of
NO3 and SO42- are 10- and 32-fold
higher today, respectively. The estimated average ANC for 60 streams in
the Shenandoah Case Study Area is 57.9 [mu]eq/L (4.5
[mu]eq/L). Fifty-five percent of all monitored streams in the
Shenandoah Case Study Area have a current risk of Elevated, Severe, or
Acute. Of the 55 percent, 18 percent experience episodic acidification,
and 18 percent are chronically acidic today.
    (4) Based on the steady-state critical load model for the year
2002, 52 percent, 72 percent, 85 percent and 93 percent of 60 modeled
streams received combined total sulfur and nitrogen deposition that
exceeded critical loads corresponding to ANC limits of 0, 20, 50, and
100 [mu]eq/L respectively.
    (5) Based on a deposition scenario that maintains current emission
levels to 2020 and 2050, the simulation forecast indicates that a large
number of streams would still have Elevated to Acute problems with
acidity.
c. Deposition-Related Terrestrial Acidification
    The role of terrestrial acidification was examined in the REA using
a critical load analysis for sugar maple and red spruce forests in the
eastern U.S. by using the BC/Al ratio in acidified forest soils as an
indicator to assess the impact of nitrogen and sulfur deposition on
tree health. These are the two most commonly studied species in North
America for impacts of acidification. At a BC/Al ratio of 1.2, red
spruce growth can be reduced by 20 percent. Sugar maple growth can be
reduced by 20 percent at a BC/Al ratio of 0.6. Key findings of the case
study are summarized below.
    (1) Case study results suggest that the health of at least a
portion of the sugar maple and red spruce growing in the U.S. may have
been compromised with acidifying total nitrogen and sulfur deposition
in 2002. The 2002 CMAQ/NADP total nitrogen and sulfur deposition levels
exceeded three selected critical loads in 3 percent to 75 percent of
all sugar maple plots across 24 states. The three critical loads ranged
from 6,008 to 107 eq/ha/yr for the BC/Al ratios of 0.6, 1.2, and 10.0
(increasing levels of tree protection). The 2002 CMAQ/NADP total
nitrogen and sulfur deposition levels exceeded three selected critical
loads in 3 percent to 36 percent of all red spruce plots across eight
states. The three critical loads

ranged from 4,278 to 180 eq/ha/yr for the Bc/Al ratios of 0.6, 1.2, and
10.0 (increasing levels of tree protection).
    (2) The SMB model assumptions made for base cation weathering (Bcw)
and forest soil ANC input parameters are the main sources of
uncertainty since these parameters are rarely measured and require
researchers to use default values.
    (3) The pattern of case study results suggests that nitrogen and
sulfur acidifying deposition in the sugar maple and red spruce forest
areas studied were similar in magnitude to the critical loads for those
areas and both ecosystems are likely to be sensitive to any future
changes in the levels of deposition.
d. Deposition-Related Aquatic Nutrient Enrichment
    The role of nitrogen deposition in two main stem rivers feeding
their respective estuaries was analyzed in the REA to determine if
decreases in deposition could influence the risk of eutrophication as
predicted using the ASSETS EI scoring system in tandem with SPARROW
modeling. This modeling approach provides a transferrable,
intermediate-level analysis of the linkages between atmospheric
deposition and receiving waters, while providing results on which
conclusions could be drawn. A summary of findings follows:
    (1) The 2002 CMAQ/NADP results showed that an estimated 40,770,000
kilograms (kg) of total nitrogen was deposited in the Potomac River
watershed. The SPARROW modeling predicted that 7,380,000 kg N/yr of the
deposited nitrogen reached the estuary (20 percent of the total load to
the estuary). The overall ASSETS EI for the Potomac River and Potomac
Estuary was Bad (based on all sources of N).
    (2) To improve the Potomac River and Potomac Estuary ASSETS EI
score from Bad to Poor, a decrease of at least 78 percent in the 2002
total nitrogen atmospheric deposition load to the watershed would be
required.
    (3) The 2002 CMAQ/NADP results showed that an estimated 18,340,000
kg of total nitrogen was deposited in the Neuse River watershed. The
SPARROW modeling predicted that 1,150,000 kg N/yr of the deposited
nitrogen reached the estuary (26 percent of the total load to the
estuary). The overall ASSETS EI for the Neuse River/Neuse River Estuary
was Bad.
    (4) It was found that the Neuse River/Neuse River Estuary ASSETS EI
score could not be improved from Bad to Poor with decreases only in the
2002 atmospheric deposition load to the watershed. Additional
reductions would be required from other nitrogen sources within the
watershed.
    The small effect of decreasing atmospheric deposition in the Neuse
River watershed is because the other nitrogen sources within the
watershed are more influential than atmospheric deposition in affecting
the total nitrogen loadings to the Neuse River Estuary, as estimated
with the SPARROW model. A water body's response to nutrient loading
depends on the magnitude (e.g., agricultural sources have a higher
influence in the Neuse than in the Potomac), spatial distribution, and
other characteristics of the sources within the watershed; therefore a
reduction in nitrogen deposition does not always produce a linear
response in reduced load to the estuary, as demonstrated by these two
case studies.
e. Deposition-Related Terrestrial Nutrient Enrichment
    California CSS and MCF communities were the focus of the
Terrestrial Nutrient Enrichment Case Studies of the REA. Geographic
information systems analysis supported a qualitative review of past
field research to identify ecological benchmarks associated with CSS
and mycorrhizal communities, as well as MCF's nutrient-sensitive
acidophyte lichen communities, fine-root biomass in Ponderosa pine and
leached nitrate in receiving waters. These benchmarks, ranging from 3.1
to 17 kg N/ha/yr, were compared to 2002 CMAQ/NADP data to discern any
associations between atmospheric deposition and changing communities.
Evidence supports the finding that nitrogen alters CSS and MCF. Key
findings include the following:
    (1) The 2002 CMAQ/NADP nitrogen deposition data show that the 3.3
kg N/ha/yr benchmark has been exceeded in more than 93 percent of CSS
areas (654,048 ha). This suggests that such deposition is a driving
force in the degradation of CSS communities. One potentially
confounding factor is the role of fire. Although CSS decline has been
observed in the absence of fire, the contributions of deposition and
fire to the CSS decline require further research. The CSS is fragmented
into many small parcels, and the 2002 CMAQ/NADP 12-km grid data are not
fine enough to fully validate the relationship between CSS
distribution, nitrogen deposition, and fire.
    (2) The 2002 CMAQ/NADP nitrogen deposition data exceeds the 3.1 kg
N/ha/yr benchmark in more than 38% (1,099,133 ha) of MCF areas, and
nitrate leaching has been observed in surface waters. Ozone effects
confound nitrogen effects on MCF acidophyte lichen, and the
interrelationship between fire and nitrogen cycling requires additional
research.
f. Additional Effects
    Ecological effects have also been documented across the U.S. where
elevated nitrogen deposition has been observed, including the eastern
slope of the Rocky Mountains where shifts in dominant algal species in
alpine lakes have occurred where wet nitrogen deposition was only about
1.5 kg N/ha/yr. High alpine terrestrial communities have a low capacity
to sequester nitrogen deposition, and monitored deposition exceeding 3
to 4 kg N/ha/yr could lead to community-level changes in plant species,
lichens and mycorrhizae.
    Additional welfare effects are documented, but examined less
extensively, in the REA. These effects include qualitative discussions
related to visibility and materials damage, such as corrosion, erosion,
and soiling of paint and buildings which are being addressed in the PM
NAAQS review currently underway. A discussion of the causal
relationship between sulfur deposition (as sulfate,
SO42-) and increased mercury methylation in
wetlands and aquatic environments is also included in the REA. On this
subject the REA concludes that decreases in SO42-
deposition will likely result in decreases in MeHg concentration;
however, spatial and biogeochemical variations nationally hinder
establishing large scale dose-response relationships.
    Several additional issues concerning oxides of nitrogen were
addressed in the REA. Consideration was also given to N2O, a
potent GHG. The REA concluded that it is most appropriate to analyze
the role of N2O in the context of all of the GHGs rather
than as part of the REA for this review. The REA considered nitrogen
deposition and its correlation with the rate of photosynthesis and net
primary productivity. Nitrogen addition ranging from 15.4 to 300 kg N/
ha/yr is documented as increasing wetland N2O production by
an average of 207 percent across all ecosystems. Nitrogen addition
ranging from 30 to 240 kg N/ha/yr increased CH4 emissions by
115 percent, averaged across all ecosystems, and methane uptake was
reduced by 38 percent averaged across all ecosystems when nitrogen
addition ranged from 10 to 560 kg N/ha/yr, but reductions were only
significant for coniferous and deciduous forests. The heterogeneity of
ecosystems across the U.S., however, introduces variations into dose-
response relationships.

    The phytotoxic effects of oxides of nitrogen and sulfur on
vegetation were also briefly discussed in the REA which concluded that
since a unique secondary NAAQS exists for SO2, and
concentrations of nitric oxide (NO), NO2 and PAN are rarely
high enough to have phytotoxic effects on vegetation, further
assessment was not warranted at this time.
3. Conclusions on Effects
    For aquatic and terrestrial acidification effects, a similar
conceptual approach was used (critical loads) to evaluate the impacts
of multiple pollutants on an ecological endpoint, whereas the
approaches used for aquatic and terrestrial nutrient enrichment were
fundamentally distinct. Although the ecological indicators for aquatic
and terrestrial acidification (i.e., ANC and BC/Al) are very different,
both ecological indicators are well-correlated with effects such as
reduced biodiversity and growth. While aquatic acidification is clearly
the targeted effect area with the highest level of confidence, the
relationship between atmospheric deposition and an ecological indicator
is also quite strong for terrestrial acidification. The main drawback
with the understanding of terrestrial acidification is that the data
are based on laboratory responses rather than field measurements. Other
stressors that are present in the field but that are not present in the
laboratory may confound this relationship.
    For nutrient enrichment effects, the REA utilized different types
of indicators for aquatic and terrestrial effects to assess both the
likelihood of adverse effects to ecosystems and the relationship
between adverse effects and atmospheric sources of oxides of nitrogen.
The ecological indicator chosen for aquatic nutrient enrichment, the
ASSETS EI, seems to be inadequate to relate atmospheric deposition to
the targeted ecological effect, likely due to the many other
confounding factors. Further, there is far less confidence associated
with the understanding of aquatic nutrient enrichment because of the
large contributions from non-atmospheric sources of nitrogen and the
influence of both oxidized and reduced forms of nitrogen, particularly
in large watersheds and coastal areas. However, a strong relationship
exists between atmospheric deposition of nitrogen and ecological
effects in high alpine lakes in the Rocky Mountains because atmospheric
deposition is the only source of nitrogen to these systems. There is
also a strong weight-of-evidence regarding the relationships between
ecological effects attributable to terrestrial nitrogen nutrient
enrichment; however, ozone and climate change may be confounding
factors. In addition, the response for other species or species in
other regions of the U.S. has not been quantified.

C. Adversity of Effects to Public Welfare

    Characterizing a known or anticipated adverse effect to public
welfare is an important component of developing any secondary NAAQS.
According to the CAA, welfare effects include: ``Effects on soils,
water, crops, vegetation, manmade materials, animals, wildlife,
weather, visibility, and climate, damage to and deterioration of
property, and hazards to transportation, as well as effect on economic
values and on personal comfort and well-being, whether caused by
transformation, conversion, or combination with other air pollutants''
(CAA, Section 302(h)). While the text above lists a number of welfare
effects, these effects do not define public welfare in and of
themselves.
    Although there is no specific definition of adversity to public
welfare, the paradigm of linking adversity to public welfare to
disruptions in ecosystem structure and function has been used broadly
by EPA to categorize effects of pollutants from the cellular to the
ecosystem level. An evaluation of adversity to public welfare might
consider the likelihood, type, magnitude, and spatial scale of the
effect as well as the potential for recovery and any uncertainties
relating to these considerations.
    Similar concepts were used in past reviews of secondary NAAQS for
ozone and PM (relating to visibility), as well as in initial reviews of
effects from lead deposition. Because oxides of nitrogen and sulfur are
deposited from ambient sources into ecosystems where they affect
changes to organisms, populations and ecosystems, the concept of
adversity to public welfare as a result of alterations in structure and
function of ecosystems is an appropriate consideration for this review.
    Based on information provided in the PA, the following section
discusses how ecological effects from deposition of oxides of nitrogen
and sulfur relate to adversity to public welfare. In the PA, public
welfare was discussed in terms of loss of ecosystem services (defined
below), which in some cases can be monetized. Each of the four main
effect areas (aquatic and terrestrial acidification and aquatic and
terrestrial nutrient over-enrichment) are discussed including current
ecological effects and associated ecosystem services.
1. Ecosystem Services
    The PA defines ecosystem services as the benefits individuals and
organizations obtain from ecosystems. Ecosystem services can be
classified as provisioning (food and water), regulating (control of
climate and disease), cultural (recreational, existence, spiritual,
educational), and supporting (nutrient cycling). Conceptually, changes
in ecosystem services may be used to aid in characterizing a known or
anticipated adverse effect to public welfare. In the REA and PA
ecosystem services are discussed as a method of assessing the magnitude
and significance to the public of resources affected by ambient
concentrations of oxides of nitrogen and sulfur and deposition in
sensitive ecosystems.
    The EPA has in previous NAAQS reviews defined ecological goods and
services for the purposes of a Regulatory Impact Analysis as the
``outputs of ecological functions or processes that directly or
indirectly contribute to social welfare or have the potential to do so
in the future. Some outputs may be bought and sold, but most are not
marketed.'' It is especially important to acknowledge that it is
difficult to measure and/or monetize the goods and services supplied by
ecosystems. It can be informative in characterizing adversity to public
welfare to attempt to place an economic valuation on the set of goods
and services that have been identified with respect to a change in
policy; however it must be noted that this valuation will be incomplete
and illustrative only.
    Knowledge about the relationships linking ambient concentrations
and ecosystem services is considered in the PA as one method by which
to inform a policy judgment on a known or anticipated adverse public
welfare effect. For example, a change in an ecosystem structure and
process, such as foliar injury, would be classified as an ecological
effect, with the associated changes in ecosystem services, such as
primary productivity, food availability, forest products, and
aesthetics (e.g., scenic viewing), classified as public welfare
effects. Additionally, changes in biodiversity would be classified as
an ecological effect, and the associated changes in ecosystem
services--productivity, existence (nonuse) value, recreational viewing
and aesthetics--would also be classified as public welfare effects.
    As described in chapters 4 and 5 of the REA, case study analyses
were performed that link deposition in sensitive ecosystems to changes
in a

given ecological indicator (e.g., for aquatic acidification, to changes
in ANC) and then to changes in ecosystems. Appendix 8 of the REA links
the changes in ecosystems to the services they provide (e.g., fish
species richness and its influence on recreational fishing). To the
extent possible for each targeted effect area, the REA linked ambient
concentrations of nitrogen and sulfur (i.e., ambient air quality
indicators) to deposition in sensitive ecosystems (i.e., exposure
pathways), and then to system response as measured by a given
ecological indicator (e.g., lake and stream acidification as measured
by ANC). The ecological effect (e.g., changes in fish species richness)
was then, where possible, associated with changes in ecosystem services
and the corresponding public welfare effects (e.g., recreational
fishing).
2. Effects on Ecosystem Services
    The process used to link ecological indicators to ecosystem
services is discussed extensively in appendix 8 of the REA. In brief,
for each case study area assessed, the ecological indicators are linked
to an ecological response that is subsequently linked to associated
services to the extent possible. For example, in the case study for
aquatic acidification the chosen ecological indicator is ANC which can
be linked to the ecosystem service of recreational fishing. Although
recreational fishing losses are the only service effects that can be
independently quantified or monetized at this time, there are numerous
other ecosystem services that may be related to the ecological effects
of acidification.
    While aquatic acidification is the focus of this proposed standard,
the other effect areas were also analyzed in the REA and these
ecosystems are being harmed by nitrogen and sulfur deposition and will
obtain some measure of protection with any decrease in that deposition
regardless of the reason for the decrease. The following summarizes the
current levels of specific ecosystem services for aquatic and
terrestrial acidification and aquatic and terrestrial nutrient over-
enrichment and attempts to quantify and when possible monetize the harm
to public welfare, as represented by ecosystem services, due to
nitrogen and sulfur deposition.
a. Aquatic Acidification
    Acidification of aquatic ecosystems primarily affects the ecosystem
services that are derived from the fish and other aquatic life found in
surface waters. In the northeastern United States, the surface waters
affected by acidification are not a major source of commercially raised
or caught fish; however, they are a source of food for some
recreational and subsistence fishers and for other consumers. Although
data and models are available for examining the effects on recreational
fishing, relatively little data are available for measuring the effects
on subsistence and other consumers. Inland waters also provide
aesthetic and educational services along with non-use services, such as
existence value (protection and preservation with no expectation of
direct use). In general, inland surface waters such as lakes, rivers,
and streams also provide a number of regulating services, playing a
role in hydrological regimes and climate regulation. There is little
evidence that acidification of freshwaters in the northeastern U.S. has
significantly degraded these specific services; however, freshwater
ecosystems also provide biological control services by providing
environments that sustain delicate aquatic food chains. The toxic
effects of acidification on fish and other aquatic life impair these
services by disrupting the trophic structure of surface waters.
Although it is difficult to quantify these services and how they are
affected by acidification, it is worth noting that some of these
services may be captured through measures of provisioning and cultural
services. For example, these biological control services may serve as
``intermediate'' inputs that support the production of ``final''
recreational fishing and other cultural services.
    As summarized in Chapter 4 of the PA, recent studies indicate that
acidification of lakes and streams can result in significant loss in
economic value. For example, data indicate that more than 9 percent of
adults in the northeastern part of the country participate annually in
freshwater fishing yielding 140 million freshwater fishing days. Each
fishing day has an estimated average value per day of $35. Therefore,
the implied total annual value of freshwater fishing in the
northeastern U.S. was $5 billion in 2006. Embedded in these numbers is
a degree of harm to recreational fishing services due to acidification
that has occurred over time. These harms have not been quantified on a
regional scale; however, a case study was conducted in the Adirondacks
area (US EPA, 2011, section 4.4.2).
    In the Adirondacks case study, estimates of changes in recreational
fishing services were determined, as well as changes more broadly in
``cultural'' ecosystem services (including recreational, aesthetic, and
nonuse services). First, the MAGIC model (US EPA, 2009, Appendix 8 and
section 2.2) was applied to 44 lakes to predict what ANC levels would
be under both ``business as usual'' conditions (i.e., allowing for some
decline in deposition due to existing regulations) and pre-emission
(i.e., background) conditions. Second, to estimate the recreational
fishing impacts of aquatic acidification in these lakes, an existing
model of recreational fishing demand and site choice was applied. This
model predicts how recreational fishing patterns in the Adirondacks
would differ and how much higher the average annual value of
recreational fishing services would be for New York residents if lake
ANC levels corresponded to background (rather than business as usual)
conditions. To estimate impacts on a broader category of cultural (and
some provisioning) ecosystem services, results from the Banzhaf et al
(2006) valuation survey of New York residents were adapted and applied
to this context. The survey used a contingent valuation approach to
estimate the average annual household willingness to pay (WTP) for
future reductions in the percent of Adirondack lakes impaired by
acidification. The focus of the survey was on impacts on aquatic
resources. Pretesting of the survey indicated that respondents
nonetheless tended to assume that benefits would occur in the condition
of birds and forests as well as in recreational fishing.
    By extrapolating the 44 lake Adirondack case study to all 3,000
Adirondack lakes and by applying the WTP survey results to all New York
residents, the study estimated aggregated benefits between $300 and
$800 million annually for the equivalent of improving lakes in the
Adirondacks region to an ANC level of 50 [mu]eq/L. The REA estimated 44
percent of the Adirondack lakes currently fall below an ANC of 50
[mu]eq/L. Several states have set goals for improving the acid status
of lakes and streams, generally targeting ANC in the range of 50 to 60
[mu]eq/L, and have engaged in costly activities to decrease
acidification.
    These results imply significant value to the public in addition to
those derived from recreational fishing services. Note that the results
are only applicable to improvements in the Adirondacks valued by
residents of New York. If similar benefits exist in other acid-impacted
areas, benefits for the nation as a whole could be substantial. The
analysis provides results on only a subset of the impacts of
acidification on ecosystem services and suggests that the

overall impact on these services is likely to be substantial.
b. Terrestrial Acidification
    Chapters 4.4.3 and 4.4.4 of the PA review several economic studies
of areas sensitive to terrestrial acidification. Forests in the
northeastern U.S. provide several important and valuable provisioning
ecosystem services, which are reflected in the production and sales of
tree products. Sugar maples are a particularly important commercial
hardwood tree species in the United States, producing timber and maple
syrup that provide hundreds of millions of dollars in economic value
annually. Red spruce is also used in a variety of wood products and
provides up to $100 million in economic value annually. Although the
data do not exist to directly link acidification damages to economic
values of lost recreational ecosystem services in forests, these
resources are valuable to the public. A recent study, reviewed in the
PA, suggests that the total annual value of recreational off-road
driving was more than $9 billion and the value of hunting and wildlife
viewing was more than $4 billion each in the northeastern States. The
EPA is not able to quantify at this time the specific effects on these
values of acid deposition, or of any specific reductions in deposition,
relative to the effects of many other factors that may affect them.
c. Nutrient Enrichment
    Chapters 4.4.5 and 4.4.6 of the PA summarize economic studies of
east coast estuaries affected by nutrient over-enrichment or
eutrophication. Estuaries in the eastern United States are important
for fish and shellfish production. The estuaries are capable of
supporting large stocks of resident commercial species, and they serve
as the breeding grounds and interim habitat for several migratory
species. To provide an indication of the magnitude of provisioning
services associated with coastal fisheries, from 2005 to 2007, the
average value of total catch was $1.5 billion per year in 15 East Coast
states. Estuaries also provide an important and substantial variety of
cultural ecosystem services, including water-based recreational and
aesthetic services. For example, data indicate that 4.8 percent of the
population in coastal states from North Carolina to Massachusetts
participated in saltwater fishing, with a total of 26 million saltwater
fishing days in 2006. Based on estimates in the PA, total recreational
value from these saltwater fishing days was approximately $1.3 billion.
Recreational participation estimates for 1999-2000 showed almost 6
million individuals participated in motorboating in coastal states from
North Carolina to Massachusetts. The aggregate value of these coastal
motorboating outings was $2 billion per year. EPA is not able to
quantify at this time the specific effects on these values of nitrogen
deposition, or of any specific reductions in deposition, relative to
the effects of many other factors that may affect them.
    Terrestrial ecosystems can also suffer from nutrient over-
enrichment. Each ecosystem is different in its composition of species
and nutrient requirements. Changes to individual ecosystems from
changes in nitrogen deposition can be hard to assess economically.
Relative recreational values are often determined by public use
information. Chapter 4.4.7 of the PA reviewed studies related to park
use in California. Data from California State Parks indicate that in
2002, 68.7 percent of adult residents participated in trail hiking for
an average of 24.1 days per year. The analyses in the PA indicate that
the aggregate annual benefit for California residents from trail hiking
in 2007 was $11.59 billion. EPA is not able to quantify at this time
the specific effects on these values of nitrogen deposition, or of any
specific reductions in deposition, relative to the effects of many
other factors that may affect them.
    The PA also identified fire regulation as a service that could be
affected by nutrient over-enrichment of the CSS and MCF ecosystems by
encouraging growth of more flammable grasses, increasing fuel loads,
and altering the fire cycle. Over the 5-year period from 2004 to 2008,
Southern California experienced, on average, over 4,000 fires per year,
burning, on average, over 400,000 acres per year. It is not possible at
this time to quantify the contribution of nitrogen deposition, among
many other factors, to increased fire risk.
3. Summary
    Adversity to public welfare can be understood by looking at how
deposition of oxides of nitrogen and sulfur affect the ecological
functions of an ecosystem (see II.A.), and then understanding the
ecosystem services that are degraded. The monetized value of the
ecosystem services provided by ecosystems that are sensitive to
deposition of oxides of nitrogen and sulfur are in the billions of
dollars each year, though it is not possible to quantify or monetize at
this time the effects on these values of nitrogen and sulfur deposition
or of any changes in deposition that may result from new secondary
standards. Many lakes and streams are known to be degraded by acidic
deposition which affects recreational fishing and tourism. Forest
growth is likely suffering from acidic deposition in sensitive areas
affecting red spruce and sugar maple timber production, sugar maple
syrup production, hiking, aesthetic enjoyment and tourism. Nitrogen
deposition contributes significantly to eutrophication in many
estuaries affecting fish production, swimming, boating, aesthetic
enjoyment and tourism. Ecosystem services are likely affected by
nutrient enrichment in many natural and scenic terrestrial areas,
affecting biodiversity, including habitat for rare and endangered
species, fire control, hiking, aesthetic enjoyment and tourism.

D. Adequacy of the Current Standards

    An important issue to be addressed in the current review of the
secondary standards for oxides of nitrogen and sulfur is whether, in
view of the scientific evidence reflected in the ISA, additional
information on exposure and risk discussed in the REA, and conclusions
drawn from the PA, the existing standards provide adequate protection.
The Administrator therefore, has considered the extent to which the
current standards are adequate for the protection of public welfare.
Having reached the general conclusion that aquatic and terrestrial
ecosystems can be degraded by deposition of oxides of nitrogen and
sulfur, it is then necessary to first evaluate the appropriateness (in
terms of form and structure) of the current standards to address the
ecological effects of oxides of nitrogen and sulfur as well as the
adequacy of the current secondary standards for oxides of nitrogen and
sulfur to provide requisite protection by considering to what degree
risks to sensitive ecosystems would be expected to occur in areas that
meet the current standards. Conclusions regarding the adequacy of the
current standards are based on the available ecological effects,
exposure and risk-based evidence. In evaluating the strength of this
information, EPA has taken into account the uncertainties and
limitations in the scientific evidence. This section addresses the
adequacy of the current standards to protect against direct exposure
effects on plants from oxides of nitrogen and sulfur, the
appropriateness of the current structure of the standards to address
deposition-related effects of oxides of nitrogen and sulfur on
sensitive ecosystems and finally, the adequacy of such standards to
protect against adverse effects related to the deposition of oxides of
nitrogen and sulfur.

1. Adequacy of the Current Standards for Direct Effects
    The current secondary oxides of nitrogen and sulfur standards are
intended to protect against adverse effects to public welfare. For
oxides of nitrogen, the current secondary standard was set identical to
the primary standard,\3\ i.e., an annual standard set for
NO2 to protect against adverse effects on vegetation from
direct exposure to ambient oxides of nitrogen. For oxides of sulfur,
the current secondary standard is a 3-hour standard intended to provide
protection for plants from the direct foliar damage associated with
atmospheric concentrations of SO2. It is appropriate to
consider whether the current standards are adequate to protect against
the direct effects on vegetation resulting from ambient NO2
and SO2 which were the basis for the current secondary
standards. The ISA concluded that there was sufficient evidence to
infer a causal relationship between exposure to SO2, NO,
NO2 and PAN and injury to vegetation. Additional research on
acute foliar injury has been limited and there is no evidence to
suggest foliar injury below the levels of the current secondary
standards for oxides of nitrogen and sulfur. There is sufficient
evidence to suggest that the levels of the current standards are likely
adequate to protect against direct phytotoxic effects.
---------------------------------------------------------------------------

    \3\ The current primary NO2 standard has recently
been changed to the 3-year average of the 98th percentile of the
annual distribution of the 1 hour daily maximum of the concentration
of NO2. The current secondary standard remains as it was
set in 1971.
---------------------------------------------------------------------------

2. Appropriateness and Adequacy of the Current Standards for
Deposition-Related Effects
    This section addresses two concepts necessary to evaluate the
current standards in the context of deposition related effects. First,
appropriateness of the current standards is considered with regard to
indicator, form, level and averaging time. This discussion centers
around the ability of the current standards to evaluate and provide
protection against deposition related effects that vary spatially and
temporally. It includes particular emphasis on the indicators and forms
of the current standards and the degree to which they are ecologically
relevant with regard to deposition related effects. Second, this
section evaluates the current standards in terms of adequacy of
protection.
a. Appropriateness
    The ISA has established that the major effects of concern for this
review of the oxides of nitrogen and sulfur standards are associated
with deposition of nitrogen and sulfur caused by atmospheric
concentrations of oxides of nitrogen and sulfur. The current standards
are not directed toward depositional effects, and none of the elements
of the current NAAQS--indicator, form, averaging time, and level--are
suited for addressing the effects of nitrogen and sulfur deposition.
    Five issues arise that call into question the ecological relevance
of the structure of the current secondary standards for oxides of
nitrogen and sulfur.
    (1) The current SO2 secondary standard (0.5 ppm
SO2 over a 3-hour average) does not utilize an exposure
period that is relevant for ecosystem impacts. The majority of
deposition related impacts are associated with depositional loads that
occur over periods of months to years. This differs significantly from
exposures associated with hourly concentrations of SO2 as
measured by the current secondary standard. By addressing short-term
concentrations, the current SO2 secondary standard, while
protective against direct foliar effects from gaseous oxides of sulfur,
does not take into account the findings of effects in the ISA, which
notes the relationship between annual deposition of sulfur and
acidification effects which are likely to be more severe and widespread
than phytotoxic effects under current ambient conditions, and include
effects from long term deposition as well as short term. Acidification
is a process that occurs over time because the ability of an aquatic
system to counteract acidic inputs is reduced as natural buffers are
used more rapidly than they can be replaced through geologic
weathering. The relevant period of exposure for ecosystems is,
therefore, not the exposures captured in the short averaging time of
the current SO2 secondary standard. The current secondary
standard for oxides of nitrogen is an annual standard (0.053 ppm
averaged over 1 year) and as such is more ecologically relevant.
    (2) Current standards do not utilize appropriate atmospheric
indicators. Nitrogen dioxide and SO2 are used as the
component of oxides of nitrogen and sulfur that are measured, but they
do not provide a complete link to the direct effects on ecosystems from
deposition of oxides of nitrogen and sulfur as they do not capture all
relevant chemical species of oxidized nitrogen and oxidized sulfur that
contribute to deposition. The ISA provides evidence that deposition
related effects are linked with total nitrogen and total sulfur
deposition, and thus all forms of oxidized nitrogen and oxidized sulfur
that are deposited will contribute to effects on ecosystems. Thus, by
using atmospheric NO2 and SO2 concentrations as
indicators, the current standards address only a fraction of total
atmospheric oxides of nitrogen and sulfur, and do not take into account
the effects from deposition of total atmospheric oxides of nitrogen and
sulfur. This suggests that more comprehensive atmospheric indicators
should be considered in designing ecologically relevant standards.
    (3) Current standards reflect separate assessments of the two
individual pollutants, NO2 and SO2, rather than
assessing the joint impacts of deposition to ecosystems. Recognizing
the role that each pollutant plays in jointly affecting ecosystem
indicators, functions, and services is vital to developing a meaningful
standard. The clearest example of this interaction is in assessment of
the impacts of acidifying deposition on aquatic ecosystems.
Acidification in an aquatic ecosystem depends on the total acidifying
potential of the deposition of both nitrogen and sulfur from both
atmospheric deposition of oxides of nitrogen and sulfur as well as the
inputs from other sources of nitrogen and sulfur such as reduced
nitrogen and non-atmospheric sources. It is the joint impact of the two
pollutants that determines the ultimate effect on organisms within the
ecosystem, and critical ecosystem functions such as habitat provision
and biodiversity. Standards that are set independently are less able to
account for the contribution of the other pollutant. This suggests that
interactions between oxides of nitrogen and oxides of sulfur should be
a critical element of the conceptual framework for ecologically
relevant standards. There are also important interactions between
oxides of nitrogen and sulfur and reduced forms of nitrogen, which also
contribute to acidification and nutrient enrichment. It is important
that the structure of the standards address the role of reduced
nitrogen in determining the ecological effects resulting from
deposition of atmospheric oxides of nitrogen and sulfur. Consideration
will also have to be given to total loadings as ecosystems respond to
all sources of nitrogen and sulfur.
    (4) Current standards do not take into account variability in
ecosystem sensitivity. Ecosystems are not uniformly distributed either
spatially or temporally in their sensitivity to oxides of nitrogen and
sulfur. Therefore, failure to account for the major determinants of
variability, including geological and soil

characteristics related to the sensitivity to acidification or nutrient
enrichment as well as atmospheric and landscape characteristics that
govern rates of deposition, may lead to standards that do not provide
requisite levels of protection across ecosystems. The current
structures of the standards do not address the complexities in the
responses of ecosystems to deposition of oxides of nitrogen and sulfur.
Ecosystems contain complex groupings of organisms that respond in
various ways to the alterations of soil and water that result from
deposition of nitrogen and sulfur compounds. Different ecosystems
therefore respond in different ways depending on a multitude of factors
that control how deposition is integrated into the system. For example,
the same levels of deposition falling on limestone dominated soils have
a very different effect from those falling on shallow glaciated soils
underlain with granite. One system may over time display no obvious
detriment while the other may experience a catastrophic loss in fish
communities. This degree of sensitivity is a function of many
atmospheric factors that control rates of deposition as well as
ecological factors that control how an ecosystem responds to that
deposition. The current standards do not take into account spatial and
seasonal variations, not only in depositional loadings, but also in
sensitivity of ecosystems exposed to those loadings. Based on the
discussion summarized above, the PA concludes that the current
secondary standards for oxides of nitrogen and oxides of sulfur are not
ecologically relevant in terms of averaging time, form, level or
indicator.
b. Adequacy of Protection
    As described in the PA, ambient conditions in 2005 indicate that
the current SO2 and NO2 secondary standards were
not exceeded at that time (US EPA, 2011, Figures 6-1 and 6-2) in
locations where negative ecological effects have been observed. In many
locations, SO2 and NO2 concentrations are
substantially below the levels of the secondary standards. This pattern
suggests that levels of deposition and any negative effects on
ecosystems due to deposition of oxides of nitrogen and sulfur under
recent conditions are occurring even though areas meet or are below
current standards. In addition, based on conclusions in the REA, these
levels will not decline in the future to levels below which it is
reasonable to anticipate effects.
    In determining the adequacy of the current secondary standards for
oxides of nitrogen and sulfur the PA considered the extent to which
ambient deposition contributes to loadings in ecosystems. Since the
last review of the secondary standard for oxides of nitrogen, a great
deal of information on the contribution of atmospheric deposition
associated with ambient oxides of nitrogen has become available. The
REA presents a thorough assessment of the contribution of oxidized
nitrogen to nitrogen deposition throughout the U.S., and the relative
contributions of ambient oxidized and reduced forms of nitrogen. The
REA concludes that based on that analysis, ambient oxides of nitrogen
are a significant component of atmospheric nitrogen deposition, even in
areas with relatively high rates of deposition of reduced nitrogen. In
addition, atmospheric deposition of oxidized nitrogen contributes
significantly to total nitrogen loadings in nitrogen sensitive
ecosystems.
    The ISA summarizes the available studies of relative nitrogen
contribution and finds that in much of the U.S., oxides of nitrogen
contribute from 50 to 75 percent of total atmospheric deposition
relative to total reactive nitrogen, which includes oxidized and
reduced nitrogen species (US EPA, 2008, section 2.8.4). Although the
proportion of total nitrogen loadings associated with atmospheric
deposition of nitrogen varies across locations, the ISA indicates that
atmospheric nitrogen deposition is the main source of new anthropogenic
nitrogen to most headwater streams, high elevation lakes, and low-order
streams. Atmospheric nitrogen deposition contributes to the total
nitrogen load in terrestrial, wetland, freshwater and estuarine
ecosystems that receive nitrogen through multiple pathways. In several
large estuarine systems, including the Chesapeake Bay, atmospheric
deposition accounts for between 10 and 40 percent of total nitrogen
loadings (US EPA, 2008).
    Atmospheric concentrations of oxides of sulfur account for nearly
all sulfur deposition in the US. For the period 2004-2006, mean sulfur
deposition in the U.S. was greatest east of the Mississippi River with
the highest deposition amount, 21.3 kg S/ha-yr, in the Ohio River
Valley where most recording stations reported 3-year averages >10 kg S/
ha-yr. Numerous other stations in the East reported S deposition >5 kg
S/ha-yr. Total sulfur deposition in the U.S. west of the 100th meridian
was relatively low, with all recording stations reporting <2 kg S/ha-yr
and many reporting <1 kg S/ha-yr. Sulfur was primarily deposited in the
form of wet SO42- followed in decreasing order by
a smaller proportion of dry SO2 and a much smaller
proportion of deposition as dry SO42-.
    As discussed throughout the REA (US EPA, 2009 and section II.B
above), there are several key areas of risk that are associated with
ambient concentrations of oxides of nitrogen and sulfur. As noted
earlier, in previous reviews of the secondary standards for oxides of
nitrogen and sulfur, the standards were designed to protect against
direct exposure of plants to ambient concentrations of the pollutants.
A significant shift in understanding of the effects of oxides of
nitrogen and sulfur has occurred since the last reviews, reflecting the
large amount of research that has been conducted on the effects of
deposition of nitrogen and sulfur to ecosystems. The most significant
current risks of adverse effects to public welfare are those related to
deposition of oxides of nitrogen and sulfur to both terrestrial and
aquatic ecosystems. These risks fall into two categories, acidification
and nutrient enrichment, which were emphasized in the REA as most
relevant to evaluating the adequacy of the existing standards in
protecting public welfare from adverse ecological effects.
i. Aquatic Acidification
    The focus of the REA case studies was on determining whether
deposition of sulfur and oxidized nitrogen in locations where ambient
oxides of nitrogen and sulfur were at or below the current standards
was resulting in acidification and related effects, including episodic
acidification and mercury methylation. Based on the case studies
conducted for lakes in the Adirondacks and streams in Shenandoah
National Park (case studies are discussed more fully in section II.B
and US EPA, 2009), there is significant risk to acid sensitive aquatic
ecosystems at atmospheric concentrations of oxides of nitrogen and
sulfur at or below the current standards. The REA also supports
strongly a relationship between atmospheric deposition of oxides of
nitrogen and sulfur and loss of ANC in sensitive ecosystems and
indicates that ANC is an excellent indicator of aquatic acidification.
The REA also concludes that at levels of deposition associated with
oxides of nitrogen and sulfur concentrations at or below the current
standards, ANC levels are expected to be below benchmark values that
are associated with significant losses in fish species richness.
    Significant portions of the U.S. are acid sensitive, and current
deposition levels exceed those that would allow

recovery of the most acid sensitive lakes in the Adirondacks (US EPA,
2008, Executive Summary). In addition, because of past loadings, areas
of the Shenandoah are sensitive to current deposition levels (US EPA,
2008, Executive Summary). Parts of the West are naturally less
sensitive to acidification and subjected to lower deposition
(particularly SOX) levels relative to the eastern United
States, and as such, less focus in the ISA is placed on the adequacy of
the existing standards in these areas, with the exception of the
mountainous areas of the West, which experience episodic acidification
due to deposition.
    In describing the effects of acidification in the two case study
areas the REA uses the approach of describing benchmarks in terms of
ANC values. Many locations in sensitive areas of the U.S. have ANC
levels below benchmark levels for ANC classified as severe, elevated,
or moderate concern (US EPA, 2011, Figure 2-1). The average current ANC
levels across 44 lakes in the Adirondack case study area is 62.1
[mu]eq/L (moderate concern). However, 44 percent of lakes had
deposition levels exceeding the critical load for an ANC of 50 [mu]eq/L
(elevated), and 28 percent of lakes had deposition levels exceeding the
(higher) critical load for an ANC of 20 [mu]eq/L (severe) (US EPA,
2009, section <IP_ADDRESS>). This information indicates that almost half of
the 44 lakes in the Adirondacks case study area are at an elevated
concern level, and almost a third are at a severe concern level. These
levels are associated with greatly diminished fish species diversity,
and losses in the health and reproductive capacity of remaining
populations. Based on assessments of the relationship between number of
fish species and ANC level in both the Adirondacks and Shenandoah
areas, the number of fish species is decreased by over half at an ANC
level of 20 [mu]eq/L relative to an ANC level at 100 [mu]eq/L (US EPA,
2009, Figure 4.2-1). When extrapolated to the full population of lakes
in the Adirondacks area using weights based on the EMAP probability
survey (US EPA, 2009, section <IP_ADDRESS>), 36 percent of lakes exceeded
the critical load for an ANC of 50 [mu]eq/L and 13 percent of lakes
exceeded the critical load for an ANC of 20 [mu]eq/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
[mu]eq/L, indicating moderate concern. However, 85 percent of these
streams had recent deposition exceeding the critical load for an ANC of
50 [mu]eq/L, and 72 percent exceeded the critical load for an ANC of 20
[mu]eq/L. As with the Adirondacks area, this information suggests that
ANC levels may decline in the future and significant numbers of
sensitive streams in the Shenandoah area are at risk of adverse impacts
on fish populations if recent conditions persist. Many other streams in
the Shenandoah area are also likely to experience conditions of
elevated to severe concern based on the prevalence in the area of
bedrock geology associated with increased sensitivity to acidification
suggesting that effects due to stream acidification could be widespread
in the Shenandoah area (US EPA, 2009, section <IP_ADDRESS>).
    In addition to these chronic acidification effects, the ISA notes
that ``consideration of episodic acidification greatly increases the
extent and degree of estimated effects for acidifying deposition on
surface waters'' (US EPA, 2008, section <IP_ADDRESS>). Some studies show
that the number of lakes that could be classified as acid-impacted
based on episodic acidification is 2 to 3 times the number of lakes
classified as acid-impacted based on chronic ANC. These episodic
acidification events can have long term effects on fish populations (US
EPA, 2008, section <IP_ADDRESS>). Under recent conditions, episodic
acidification has been observed in locations in the eastern U.S. and in
the mountainous western U.S. (US EPA, 2008, section <IP_ADDRESS>).
    The ISA, REA and PA all conclude that the current standards are not
adequate to protect against the adverse impacts of aquatic
acidification on sensitive ecosystems. A recent survey, as reported in
the ISA, found sensitive streams in many locations in the U.S.,
including the Appalachian Mountains, the Coastal Plain, and the
Mountainous West (US EPA, 2008, section <IP_ADDRESS>). In these sensitive
areas, between 1 and 6 percent of stream kilometers are chronically
acidified. The REA further concludes that both the Adirondack and
Shenandoah case study areas are currently receiving deposition from
ambient oxides of nitrogen and sulfur in excess of their ability to
neutralize such inputs. In addition, based on the current emission
scenarios, forecast modeling out to the year 2020 as well as 2050
indicates a large number of streams in these areas will still be
adversely impacted (section II.B). Based on these considerations, the
PA concludes that the current secondary NAAQS for oxides of nitrogen
and sulfur do not provide adequate protection of sensitive ecosystems
with regard to aquatic acidification.
ii. Terrestrial Acidification
    Based on the terrestrial acidification case studies, Kane
Experimental Forest in Pennsylvania and Hubbard Brook Experimental
Forest described in section II.B) of sugar maple and red spruce
habitat, the REA concludes that there is significant risk to sensitive
terrestrial ecosystems from acidification at atmospheric concentrations
of NO2 and SO2 at or below the current standards.
The ecological indicator selected for terrestrial acidification is the
BC/Al, which has been linked to tree health and growth. The results of
the REA strongly support a relationship between atmospheric deposition
of oxides of nitrogen and sulfur and BC/Al, and that BC/Al is a good
indicator of terrestrial acidification. At levels of deposition
associated with oxides of nitrogen and sulfur concentrations at or
below the current standards, BC/Al levels are expected to be below
benchmark values that are associated with significant effects on tree
health and growth. Such degradation of terrestrial ecosystems could
affect ecosystem services such as habitat provisioning, endangered
species, goods production (timber, syrup, etc.) among others.
    Many locations in sensitive areas of the U.S. have BC/Al levels
below benchmark levels classified as providing low to intermediate
levels of protection to tree health. At a BC/Al ratio of 1.2
(intermediate level of protection), red spruce growth can be reduced by
20 percent. At a BC/Al ratio of 0.6 (low level of protection), sugar
maple growth can be decreased by 20 percent. The REA did not evaluate
broad sensitive regions. However, in the sugar maple case study area
(Kane Experimental Forest), recent deposition levels are associated
with a BC/Al ratio below 1.2, indicating between intermediate and low
level of protection, which would indicate the potential for a greater
than 20 percent reduction in growth. In the red spruce case study area
(Hubbard Brook Experimental Forest), recent deposition levels are
associated with a BC/Al ratio slightly above 1.2, indicating slightly
better than an intermediate level of protection (US EPA, 2009, section
<IP_ADDRESS>).
    Over the full range of sugar maple, 12 percent of evaluated forest
plots exceeded the critical loads for a BC/Al ratio of 1.2, and 3
percent exceeded the critical load for a BC/Al ratio of 0.6. However,
there was large variability across states. In New Jersey, 67 percent of
plots exceeded the critical load for a

BC/Al ratio of 1.2, while in several states on the outskirts of the
range for sugar maple (e.g. Arkansas, Illinois) no plots exceeded the
critical load for a BC/Al ratio of 1.2. For red spruce, overall 5
percent of plots exceeded the critical load for a BC/Al ratio of 1.2,
and 3 percent exceeded the critical load for a BC/Al ratio of 0.6. In
the major red spruce producing states (Maine, New Hampshire, and
Vermont), critical loads for a BC/Al ratio of 1.2 were exceeded in 0.5,
38, and 6 percent of plots, respectively.
    The ISA, REA and PA all conclude that the current standards are not
adequate to protect against the adverse impacts of terrestrial
acidification on sensitive ecosystems. As stated in the REA and PA, the
main drawback, with the understanding of terrestrial acidification lies
in the sparseness of available data by which we can predict critical
loads and that the data are based on laboratory responses rather than
field measurements. Other stressors that are present in the field but
that are not present in the laboratory may confound this relationship.
The REA does however, conclude that the case study results, when
extended to a 27 state region, show that nitrogen and sulfur acidifying
deposition in the sugar maple and red spruce forest areas caused the
calculated Bc/Al ratio to fall below 1.2 (the intermediate level of
protection) in 12 percent of the sugar maple plots and 5 percent of the
red spruce plots; however, results from individual states ranged from 0
to 67 percent of the plots for sugar maple and 0 to 100 percent of the
plots for red spruce.
iii. Terrestrial Nutrient Enrichment
    Nutrient enrichment effects are due to nitrogen loadings from both
atmospheric and non-atmospheric sources. Evaluation of nutrient
enrichment effects requires an understanding that nutrient inputs are
essential to ecosystem health and that specific long term levels of
nutrients in a system affect the types of species that occur over long
periods of time. Short term additions of nutrients can affect species
competition, and even small additions of nitrogen in areas that are
traditionally nutrient poor can have significant impacts on
productivity as well as species composition. Most ecosystems in the
U.S. are nitrogen-limited, so regional decreases in emissions and
deposition of airborne nitrogen compounds could lead to some decrease
in growth of the vegetation that surrounds the targeted aquatic system
but as discussed below evidence for this is mixed. Whether these
changes in plant growth are seen as beneficial or adverse will depend
on the nature of the ecosystem being assessed.
    Information on the effects of changes in nitrogen deposition on
forestlands and other terrestrial ecosystems is very limited. The
multiplicity of factors affecting forests, including other potential
stressors such as ozone, and limiting factors such as moisture and
other nutrients, confound assessments of marginal changes in any one
stressor or nutrient in forest ecosystems. The ISA notes that only a
fraction of the deposited nitrogen is taken up by the forests, most of
the nitrogen is retained in the soils (US EPA, 2008, section <IP_ADDRESS>).
In addition, the ISA indicates that forest management practices can
significantly affect the nitrogen cycling within a forest ecosystem,
and as such, the response of managed forests to nitrogen deposition
will be variable depending on the forest management practices employed
in a given forest ecosystem (US EPA, 2008, Annex C C.6.3). Increases in
the availability of nitrogen in nitrogen-limited forests via
atmospheric deposition could increase forest production over large non-
managed areas, but the evidence is mixed, with some studies showing
increased production and other showing little effect on wood production
(US EPA, 2008, section 3.3.9). Because leaching of nitrate can promote
cation losses, which in some cases create nutrient imbalances, slower
growth and lessened disease and freezing tolerances for forest trees,
the net effect of increased N on forests in the U.S. is uncertain (US
EPA, 2008, section 3.3.9).
    The scientific literature has many examples of the deleterious
effects caused by excessive nitrogen loadings to terrestrial systems.
Several studies have set benchmark values for levels of N deposition at
which scientifically adverse effects are known to occur. Large areas of
the country appear to be experiencing deposition above these
benchmarks. The ISA indicates studies that have found that at 3.1 kg N/
ha/yr, the community of lichens begins to change from acidophytic to
tolerant species; at 5.2 kg N/ha/yr, the typical dominance by
acidophytic species no longer occurs; and at 10.2 kg N/ha/yr,
acidophytic lichens are totally lost from the community. Additional
studies in the Colorado Front Range of the Rocky Mountain National Park
support these findings. These three values (3.1, 5.2, and 10.2 kg/ha/
yr) are one set of ecologically meaningful benchmarks for the mixed
conifer forest (MCF) of the pacific coast regions. Nearly all of the
known sensitive communities receive total nitrogen deposition levels
above the 3.1 N kg/ha/yr ecological benchmark according to the 12 km,
2002 CMAQ/NADP data, with the exception of the easternmost Sierra
Nevadas. The MCFs in the southern portion of the Sierra Nevada forests
and nearly all MCF communities in the San Bernardino forests receive
total nitrogen deposition levels above the 5.2 N kg/ha/yr ecological
benchmark.
    Coastal Sage Scrub communities are also known to be sensitive to
community shifts caused by excess nitrogen loadings. Studies have
investigated the amount of nitrogen utilized by healthy and degraded
CSS systems. In healthy stands, the authors estimated that 3.3 kg N/ha/
yr was used for CSS plant growth. It is assumed that 3.3 kg N/ha/yr is
near the point where nitrogen is no longer limiting in the CSS
community and above which level community changes occur, including
dominance by invasive species and loss of coastal sage scrub.
Therefore, this amount can be considered an ecological benchmark for
the CSS community. The majority of the known CSS range is currently
receiving deposition in excess of this benchmark. Thus, the REA
concludes that recent conditions where oxides of nitrogen ambient
concentrations are at or below the current oxides of nitrogen secondary
standards are not adequate to protect against anticipated adverse
impacts from N nutrient enrichment in sensitive ecosystems.
iv. Aquatic Nutrient Enrichment
    The REA aquatic nutrient enrichment case studies focused on coastal
estuaries and revealed that while current ambient loadings of
atmospheric oxides of nitrogen are contributing to the overall
depositional loading of coastal estuaries, other non-atmospheric
sources are contributing in far greater amounts in total, although
atmospheric contributions are as large as some other individual source
types. The ability of current data and models to characterize the
incremental adverse impacts of nitrogen deposition is limited, both by
the available ecological indicators, and by the inability to attribute
specific effects to atmospheric sources of nitrogen. The REA case
studies used ASSETS EI as the ecological indicator for aquatic nutrient
enrichment. This index is a six level index characterizing overall
eutrophication risk in a water body. This indicator is not sensitive to
changes in nitrogen deposition within a single level of the index. In
addition, this type of indicator does not reflect the impact of
nitrogen deposition in conjunction with other sources of nitrogen.

    Based on the above considerations, the REA concludes that the
ASSETS EI is not an appropriate ecological indicator for estuarine
aquatic eutrophication and that additional analysis is required to
develop an appropriate indicator for determining the appropriate levels
of protection from N nutrient enrichment effects in estuaries related
to deposition of oxides of nitrogen. As a result, EPA is unable to make
a determination as to the adequacy of the existing secondary oxides of
nitrogen standard in protecting public welfare from nitrogen nutrient
enrichment effects in estuarine aquatic ecosystems.
    Additionally, nitrogen deposition can alter species composition and
cause eutrophication in freshwater systems. In the Rocky Mountains, for
example, deposition loads of 1.5 to 2 kg/ha/yr which are well within
current ambient levels are known to cause changes in species
composition in diatom communities indicating impaired water quality (US
EPA, 2008, section <IP_ADDRESS>). This suggests that the existing secondary
standard for oxides of nitrogen does not protect such ecosystems and
their resulting services from impairment.
v. Other Effects
    An important consideration in looking at the effects of deposition
of oxides of sulfur in aquatic ecosystems is the potential for
production of MeHg, a neurotoxic contaminant. The production of
meaningful amounts of MeHg requires the presence of
SO42- and mercury, and where mercury is present,
increased availability of SO42- results in
increased production of MeHg. There is increasing evidence on the
relationship between sulfur deposition and increased methylation of
mercury in aquatic environments; this effect occurs only where other
factors are present at levels within a range to allow methylation. The
production of MeHg requires the presence of SO42-
and mercury, but the amount of MeHg produced varies with oxygen
content, temperature, pH and supply of labile organic carbon (US EPA,
2008, section 3.4). In watersheds where changes in sulfate deposition
did not produce an effect, one or several of those interacting factors
were not in the range required for meaningful methylation to occur (US
EPA, 2008, section 3.4). Watersheds with conditions known to be
conducive to mercury methylation can be found in the northeastern
United States and southeastern Canada (US EPA, 2009, section 6).
    With respect to sulfur deposition and mercury methylation, the
final ISA determined that ''[t]he evidence is sufficient to infer a
causal relationship between sulfur deposition and increased mercury
methylation in wetlands and aquatic environments.'' However, EPA did
not conduct a quantitative assessment of the risks associated with
increased mercury methylation under current conditions. As such, EPA is
unable to make a determination as to the adequacy of the existing
SO2 secondary standards in protecting against welfare
effects associated with increased mercury methylation.
vi. Summary of Adequacy Considerations
    In summary, the PA concludes that currently available scientific
evidence and assessments clearly call into question the adequacy of the
current standards with regard to deposition-related effects on
sensitive aquatic and terrestrial ecosystems, including acidification
and nutrient enrichment. Further, the PA recognizes that the elements
of the current standards--indicator, averaging time, level and form--
are not ecologically relevant, and are thus not appropriate for
standards designed to provide such protection. Thus, the PA concludes
that consideration should be given to establishing a new ecologically
relevant multi-pollutant, multimedia standard to provide appropriate
protection from deposition-related ecological effects of oxides of
nitrogen and sulfur on sensitive ecosystems, with a focus on protecting
against adverse effects associated with acidifying deposition in
sensitive aquatic ecosystems.
3. CASAC Views
    In a letter to the Administrator (Russell and Samet 2011a), the
CASAC Oxides of Nitrogen and Oxides of Sulfur Panel, with full
endorsement of the chartered CASAC, unanimously concluded that:

    EPA staff has demonstrated through the Integrated Science
Assessment (ISA), Risk and Exposure Characterization (REA) and the
draft PA that ambient NOX and SOX can have,
and are having, adverse environmental impacts. The Panel views that
the current NOX and SOX secondary standards
should be retained to protect against direct adverse impacts to
vegetation from exposure to gas phase exposures of these two
families of air pollutants. Further, the ISA, REA and draft PA
demonstrate that adverse impacts to aquatic ecosystems are also
occurring due to deposition of NOX and SOX.
Those impacts include acidification and undesirable levels of
nutrient enrichment in some aquatic ecosystems. The levels of the
current NOX and SOX secondary NAAQS are not
sufficient, nor the forms of those standards appropriate, to protect
against adverse depositional effects; thus a revised NAAQS is
warranted.

    In addition, with regard to the joint consideration of both oxides
of nitrogen and oxides of sulfur as well as the consideration of
deposition related effects, CASAC concluded that the PA had developed a
credible methodology for considering such effects. The Panel stated
that ``the Policy Assessment develops a framework for a multi-
pollutant, multimedia standard that is ecologically relevant and
reflects the combined impacts of these two pollutants as they deposit
to sensitive aquatic ecosystems.''
4. Administrator's Proposed Conclusions Concerning Adequacy of Current
Standard
    Based on the above considerations and taking into account CASAC
advice, the Administrator recognizes that the purpose of the secondary
standard is to protect against ``adverse'' effects resulting from
exposure to oxides of nitrogen and sulfur, discussed above in section
II.A. The Administrator also recognizes the need for conclusions as to
the adequacy of the current standards for both direct and deposition
related effects as well as conclusions as to the appropriateness and
ecological relevance of the current standards.
    In considering what constitutes an ecological effect that is also
adverse to the public welfare, the Administrator took into account the
ISA conclusions regarding the nature and strength of the effects
evidence, the risk and exposure assessment results, the degree to which
the associated uncertainties should be considered in interpreting the
results, the conclusions presented in the PA, and the views of CASAC
and members of the public. On these bases, the Administrator concludes
that the current secondary standards are adequate to protect against
direct phytotoxic effects on vegetation. Thus, the Administrator
proposes to retain the current secondary standard for oxides of
nitrogen at 53 ppb,\4\ annual average concentration, measured in the
ambient air as NO2, and the current secondary standard for
oxides of sulfur at 0.5 ppm,

3-hour average concentration, measured in the ambient air as
SO2.
---------------------------------------------------------------------------

    \4\ The annual secondary standard for oxides of nitrogen is
being specified in units of ppb to conform to the current version of
the annual primary standard, as specified in the final rule for the
most recent review of the NO2 primary NAAQS (75 FR 6531;
February 9, 2010).
---------------------------------------------------------------------------

    With regard to deposition-related effects, the Administrator has
first to consider the appropriateness of the structure of the current
standards to address ecological effects of concern. Based on the
evidence as well as considering the advice given by CASAC on this
matter, the Administrator concludes that the elements of the current
standards are not ecologically relevant and thus are not appropriate to
provide protection of ecosystems. On the subject of adequacy of
protection with regard to deposition-related effects, the Administrator
considered the full nature of ecological effects related to the
deposition of ambient oxides of nitrogen and sulfur into sensitive
ecosystems across the country. Her conclusions are based on the
evidence presented in the ISA with regard to acidification and nutrient
enrichment effects, the findings of the REA with regard to scope and
severity of the current and likely future effects of deposition, the
synthesis of both the scientific evidence and risk and exposure results
in the PA as to the adequacy of the current standards, and the advice
of both CASAC and the public. After such consideration, the
Administrator concludes that current levels of oxides of nitrogen and
sulfur are sufficient to cause acidification of both aquatic and
terrestrial ecosystems, nutrient enrichment of terrestrial ecosystems
and contribute to nutrient enrichment effects in estuaries that could
be considered adverse, and the current secondary standards do not
provide adequate protection from such effects.
    Having reached these conclusions, the Administrator determines that
it is appropriate to consider alternative standards that are
ecologically relevant. These considerations support the conclusion that
the current secondary standards is neither appropriate nor adequate to
protect against deposition related effects. The Administrator's
consideration of such alternative standards is discussed below in
Section III.

III. Rationale for Proposed Decision on Alternative Multi-Pollutant
Approach to Secondary Standards for Aquatic Acidification

    Having reached the conclusion that the current NO2 and
SO2 secondary standards are not adequate to provide
appropriate protection against deposition-related effects associated
with oxides of nitrogen and sulfur, the Administrator then considered
what new multi-pollutant standard might be appropriate, at this time,
to address such effects on public welfare. The Administrator recognizes
that the inherently complex and variable linkages between ambient
concentrations of nitrogen and sulfur oxides, the related deposited
forms of nitrogen and sulfur, and the ecological responses that are
associated with public welfare effects call for consideration of an
ecologically relevant design of a standard that reflects these
linkages. The Administrator also recognizes that characterization of
such complex and variable linkages will necessarily require
consideration of information and analyses that have important
limitations and uncertainties.
    Despite its complexity, an ecologically relevant multi-pollutant
standard to address deposition-related effects could still
appropriately be defined in terms of the same basic elements that are
used to define any NAAQS--indicator, form, averaging time, and level.
The form would incorporate additional structural elements that reflect
relevant multi-pollutant and multimedia attributes. These structural
elements include the use of an ecological indicator, tied to the
ecological effect we are focused on, and other elements that account
for ecologically relevant factors other than ambient air
concentrations. All of these elements would be needed to enable a
linkage from ambient air indicators to the ecological indicator to
define an ecologically relevant standard. As a result, such a standard
would necessarily be more complex than the NAAQS that have been set
historically to address effects associated with ambient concentrations
of a single pollutant.
    More specifically, the Administrator considered an ecologically
relevant multi-pollutant standard to address effects associated with
acidifying deposition related to ambient concentrations of oxides of
nitrogen and sulfur in sensitive aquatic ecosystems. This focus is
consistent with the information presented in the ISA, REA, and PA,
which highlighted the sufficiency of the quantity and quality of the
available evidence and assessments associated with aquatic
acidification relative to the information and assessments available for
other deposition-related effects, including terrestrial acidification
and aquatic and terrestrial nutrient enrichment. Based on its review of
these documents, CASAC agreed that aquatic acidification should be the
focus for developing a new multi-pollutant standard in this review. In
reaching conclusions about an air quality standard designed to address
deposition-related aquatic acidification effects, the Administrator
also recognizes that such a standard may also provide some degree of
protection against other deposition-related effects.
    As discussed in chapter 7 of the PA, the development of a new
multi-pollutant standard to address deposition-related aquatic
acidification effects recognizes the need for consideration of a
nationally applicable standard for protection against adverse effects
of aquatic acidification on public welfare, while recognizing the
complex and heterogeneous interactions between ambient air
concentrations of nitrogen and sulfur oxides, the related deposition of
nitrogen and sulfur, and associated ecological responses. The
development of such a standard also needs to take into account the
limitations and uncertainties in the available information and analyses
upon which characterization of such interactions are based. The
approach used in the PA also recognizes that while such a standard
would be national in scope and coverage, the effects to public welfare
from aquatic acidification will not occur to the same extent in all
locations in the U.S., given the inherent variability of the responses
of aquatic systems to the effects of acidifying deposition.
    As discussed above in section II, many locations in the U.S. are
naturally protected against acid deposition due to underlying
geological conditions. Likewise, some locations in the U.S., including
lands managed for commercial agriculture and forestry, are not likely
to be negatively impacted by current levels of nitrogen and sulfur
deposition. As a result, while a new ecologically relevant secondary
standard would apply everywhere, it would be structured to account for
differences in the sensitivity of ecosystems across the country. This
would allow for appropriate protection of sensitive aquatic ecosystems,
which are relatively pristine and wild and generally in rural areas,
and the services provided by such sensitive ecosystems, without
requiring more protection than is needed elsewhere.
    As discussed below, the multi-pollutant standard developed in the
PA would employ (1) total reactive oxidized nitrogen (NOy)
and SOX as the atmospheric ambient air indicators; (2) a
form that takes into account variable factors, such as atmospheric and
ecosystem conditions that modify the amounts of deposited nitrogen and
sulfur; the distinction between oxidized and reduced forms of nitrogen;
effects of deposited nitrogen and sulfur on aquatic ecosystems in terms
of the ecological indicator ANC; and the

representativeness of water bodies within a defined spatial area; (3) a
multi-year averaging time, and (4) a standard level defined in terms of
a single, national target ANC value that, in the context of the above
form, identifies the levels of concentrations of NOy and
SOX in the ambient air that would meet the standard. The
form of such a standard has been defined by an index, AAI, which
reflects the relationship between ambient concentrations of
NOy and SOX and aquatic acidification effects
that result from nitrogen and sulfur deposition related to these
ambient concentrations.
    In presenting the considerations associated with such an air
quality standard to address deposition-related aquatic acidification
effects, the following sections focus on each element of the standard,
including indicator (section III.A), form (section III.B), averaging
time (section III.C), and level (section III.D). Alternative
combinations of levels and forms are discussed in section III.E.
Considerations related to important uncertainties inherent in such an
approach are discussed in section III.F. Advice from CASAC on such a
new standard is presented in section III.G. The Administrator's
proposed decisions on such a new standard are presented in section
III.H.

A. Ambient Air Indicators

    In considering alternative ambient air indicators, the PA primarily
focuses on the important attribute of association. Association in a
broad sense refers to how well an ambient air indicator relates to the
ecological effects of interest by virtue of both the framework that
links the ambient indicator and effects and the empirical evidence that
quantifies the linkages. The PA also considers how measurable or
quantifiable an indicator is to enable its use as an effective
indicator of relevant ambient air concentrations.
    As discussed above in section II.C, the PA concludes that
indicators other than NO2 and SO2 should be
considered as the appropriate indicators of oxides of nitrogen and
sulfur in the ambient air for protection against the acidification
effects associated with deposition of the associated nitrogen and
sulfur. This conclusion is based on the recognition that all forms of
nitrogen and sulfur in the ambient air contribute to deposition and
resulting acidification, and as such, NO2 and SO2
are incomplete indicators. In principle, the ambient indicators should
represent the species that are associated with oxides of nitrogen and
sulfur in the ambient air and can contribute acidifying deposition.
This includes both the species of oxides of nitrogen and sulfur that
are directly emitted as well as species transformed in the atmosphere
from oxides of nitrogen and sulfur that retain the nitrogen and sulfur
atoms from directly emitted oxides of nitrogen and sulfur. All of these
compounds are associated with oxides of nitrogen and sulfur in the
ambient air and can contribute to acidifying deposition.
    The PA focuses in particular on the various compounds with nitrogen
or sulfur atoms that are associated with oxides of nitrogen and sulfur,
because the acidifying potential is specific to nitrogen and sulfur,
and not other atoms (e.g., H, C, O) whether derived from the original
source of oxides of nitrogen and sulfur emissions or from atmospheric
transformations. For example, the acidifying potential of each molecule
of NO2, NO, HNO3 or PAN is identical, as is the
potential for each molecule of SO2 or ion of particulate
sulfate, p-SO4. Each atom of sulfur affords twice the
acidifying potential of each atom of nitrogen.
1. Oxides of Sulfur
    As discussed in the PA (US EPA, 2011, section 7.1.1), oxides of
sulfur include the gases sulfur monoxide (SO), SO2, sulfur
trioxide (SO3), disulfur monoxide (S2O), and
particulate-phase sulfur compounds (referred to as SO4) that
result from gas-phase sulfur oxides interacting with particles.
However, the sum of SO2 and SO4 does represent
virtually the entire ambient air mass of sulfur that contributes to
acidification. In addition to accounting for virtually all the
potential for acidification from oxidized sulfur in the ambient air,
there are reliable methods to monitor the concentrations of
SO2 and particulate SO4. In addition, much of the
data used to develop the technical basis for the standard developed in
the PA is based on monitoring or modeling of these species.\5\ The PA
concludes that the sum of SO2 and SO4, referred
to as SOX, are appropriate ambient air indicators of oxides
of sulfur because they represent virtually all of the acidification
potential of ambient air oxides of sulfur and there are reliable
methods suitable for measuring SO2 and SO4.
---------------------------------------------------------------------------

    \5\ As discussed in chapter 2 of the PA, SO2 and
particulate SO4 are routinely measured in ambient air
monitoring networks, although only the Clean Air Status and Trends
Network (CASTNET) filter packs do not intentionally exclude particle
size fractions. The CMAQ treatment of SOX is the simple
addition of both species, which are treated explicitly in the model
formulation. All particle size fractions are included in the CMAQ
SOX estimates.
---------------------------------------------------------------------------

2. Oxides of Nitrogen
    As discussed in the PA (US EPA, 2011, section 7.1.2),
NOy, as defined in chapter 2 of the PA, incorporates
basically all of the oxidized nitrogen species that have acidifying
potential and as such, NOy should be considered as an
appropriate indicator for oxides of nitrogen. Total reactive oxidized
nitrogen is an aggregate measure of NO and NO2 and all of
the reactive oxidized products of NO and NO2. That is,
NOy is a group of nitrogen compounds in which all of the
compounds are either an oxide of nitrogen or compounds in which the
nitrogen atoms came from oxides of nitrogen. Total reactive oxidized
nitrogen is especially relevant as an ambient indicator for
acidification in that it both relates to the oxides of nitrogen in the
ambient air and also represents the acidification potential of all
oxidized nitrogen species in the ambient air, whether an oxide of
nitrogen or derived from oxides of nitrogen.
    There are currently available reliable methods of measuring
aggregate NOy. The term ``aggregate'' measure means that the
NOy, as measured, is not based on measuring each individual
species of NOy and calculating an NOy value by
summing the individual species. Rather, as described in chapter 2 of
the PA, current measurement techniques process all of the individual
NOy species to produce a single aggregate measure of all of
the nitrogen atoms associated with any NOy species.
Consequently, the NOy measurement effectively provides the
sum of all individual species, but the identity of the individual
species is lost. As discussed above, the accounting for the individual
nitrogen atoms is an accounting of the ambient air acidification
potential of oxides of nitrogen and their transformation products and
therefore the most relevant ambient indicator for aquatic acidification
effects associated with oxides of nitrogen.
    This loss of the information on individual species motivated
consideration of alternative or more narrowly defined indicators for
oxides of nitrogen in the PA. Consideration of a subset of
NOy species was based on the following reasoning. First, the
actual dry deposition of nitrogen is determined on an individual
species basis by multiplying the species concentration times a species-
specific deposition velocity and then summed to develop an estimate of
total dry deposition. Consequently, the use of individual ambient
species has the potential to be more consistent with the underlying

science of deposition and, therefore, has the potential to allow for a
more rigorous evaluation of dry deposition with specialized field
studies. In addition, there has been a suggestion of focusing only on
the most quickly depositing NOy species, such as
HNO3, as contributions from other NOy species
such as NO2 may be negligible. These alternative indicators
are discussed below.
    The PA considers the relative merits of using each individual
NOy species as part of a group of indicators. In so doing,
it was first noted that dry deposition of NOy is treated as
the sum of the deposition of each individual species in advanced
process-based air quality models like CMAQ, as described in chapter 2
of the PA. Conceptually one could extend this process-based approach by
using all NOy species individually as separate indicators
for oxides of nitrogen and requiring, for example, measurements of each
of the species, including the dominant species of HNO3,
particulate nitrate (p-NO3), true NO2, NO and
PAN. The potential attraction of using individual species would be the
reliance on actual deposition velocities. This could have more physical
meaning in comparison to a constructed model of aggregate deposition of
NOy, which is difficult to evaluate with observations
because of the assimilation of many species with disparate deposition
behavior. The PA notes that the major drawback of using individual
NOy species as the indicators is the lack of reliable
measurement techniques, especially for PAN and NO2 in rural
locations, which renders the use of virtually any individual
NOy species, except for NO and perhaps p-NO3, as
functionally inadequate from a measurement perspective.
    The PA next considered the relative merits of using a subset of
NOy species as the indicators for oxides of nitrogen, as was
discussed above for oxides of sulfur. To the extent that certain
species provide relatively minor contributions to total NOy
deposition, it may be appropriate to consider excluding them as part of
the indicator. As discussed in chapter 2 of the PA, each nitrogen
species within the array of NOy species has species-specific
dry deposition velocities. For example, the deposition velocity of
HNO3 is much greater than the velocity for NO2
and, consequently, for a similar ambient air concentration,
HNO3 contributes more deposition of acidifying nitrogen
relative to NO2. In transitioning from source-oriented urban
locations to rural environments, the ratio of the concentrations of
HNO3 and PAN to NO2 increases.
    Based on the reasoning that a larger fraction of the deposited
NOy is accounted for by total nitrate (the sum of
HNO3 and p-NO3), a surrogate for the more rapidly
depositing fraction of NOy, combined with the availability
of reliable total nitrate measurements through the CASTNET, the PA
considered using total nitrate as the indicator for oxides of nitrogen
(US EPA, 2011, appendix E). Nitrate would be expected to correlate well
with total reactive oxidized nitrogen deposition relative to
NOy (US EPA, 2011, chapter 2) despite the inherent noise
associated with variable contributions of low deposition velocity
species (e.g., NO2) that may have relatively high ambient
concentrations. However, modeling simulations suggest that
NOy may be a more robust indicator, relative to
HNO3, in terms of relating absolute changes in ambient air
concentrations to changes in nitrogen deposition driven by changes in
ambient concentrations of oxides of nitrogen (US EPA, 2011, Figure 2-
32).
    Based on the above considerations, the PA concludes that
NOy should be considered as the appropriate ambient
indicator for oxides of nitrogen based on its direct relationship to
oxides of nitrogen in the ambient air and its direct relationship to
deposition associated with aquatic acidification. Because
NOy represents all of the potentially acidifying oxidized
nitrogen species in the ambient air, it is appropriately associated
with the deposition of potentially acidifying compounds associated with
oxides of nitrogen in the ambient air. In addition, there are reliable
methods available to measure NOy. Measurement of each
individual species of NOy, or the measurement of only a
subset of species of NOy, is less appropriate because there
are not reliable measurements methods available to measure all of the
individual species of NOy and a subset of species would fail
to account for significant portions of the oxidized reactive nitrogen
that relate to acidification.\6\
---------------------------------------------------------------------------

    \6\ The PA also notes that NOy is a useful
measurement for model evaluation purposes, which is especially
important, recognizing the unique role that CMAQ plays in the
development of this standard, as described below in section III.B.
---------------------------------------------------------------------------

B. Form

    Based on the evidence of the aquatic acidification effects caused
by the deposition of NOy and SOX, the PA (US EPA,
2011, section 7.2) presents the development of a new form that is
ecologically relevant for addressing such effects. The conceptual
design for the form of such a standard includes three main components:
an ecological indicator, deposition metrics that relate to the
ecological indicator, and a function that relates ambient air
indicators to deposition metrics. Collectively, these three components
link the ecological indicator to ambient air indicators, as illustrated
above in Fig II-1.
    The simplified flow diagram in Figure II-1 compresses the various
atmospheric, biological, and geochemical processes associated with
acidifying deposition to aquatic ecosystems into a simplified
conceptual picture. The ecological indicator (left box) is related to
atmospheric deposition through biogeochemical ecosystem models (middle
box), which associate a target deposition load to a target ecological
indicator. Once a target deposition is established, associated
allowable air concentrations are determined (right box) through the
relationships between concentration and deposition that are embodied in
air quality models such as CMAQ. The following discussion describes the
development and rationale for each of these components, as well as the
integration of these components into the full expression of the form of
the standard using the concept of a national AAI that represents a
target ANC level as a function of ambient air concentrations. Spatial
aggregation issues associated with defining each of the terms of this
index are also addressed below.
    The AAI is designed to be an ecologically relevant form of the
standard that determines the levels of NOy and
SOX in the ambient air that would achieve a target ANC limit
for the U.S. The intent of the AAI is to weight atmospheric
concentrations of oxides of nitrogen and sulfur by their propensity to
contribute to acidification through deposition, given the fundamental
acidifying potential of each pollutant, and to take into account the
ecological factors that govern acid sensitivity in different
ecosystems. The index also accounts for the contribution of reduced
nitrogen to acidification. Thus, the AAI encompasses those attributes
of specific relevance to protecting ecosystems from the acidifying
potential of ambient air concentrations of NOy and
SOX.
1. Ecological Indicator
    In considering alternative ecological indicators, the PA again
primarily focuses on the attribute of association. In the case of an
ecological indicator for aquatic acidification, association refers to
the relationship between the indicator and adverse effects as discussed
in section II. Because of the conceptual structure of the form of an

AAI-based standard (Figure III-1), this particular ecological indicator
must also link up in a meaningful and quantifiable manner with
acidifying atmospheric deposition. In effect, the ecological indicator
for aquatic acidification is the bridge between biological impairment
and deposition of NOy and SOX.
    This section presents the rationale in the PA for selecting ANC as
the appropriate ecological indicator for consideration. Recognizing
that ANC is not itself the causative or toxic agent for adverse aquatic
acidification effects, the rationale for using ANC as the relevant
ecological indicator is based on the following:
    (1) The ANC is directly associated with the causative agents, pH
and dissolved Al, both through empirical evidence and mechanistic
relationships;
    (2) Empirical evidence shows very clear and strong relationships
between adverse effects and ANC;
    (3) The ANC is a more reliable indicator from a modeling
perspective, allowing use of a body of studies and technical analyses
related to ANC and acidification to inform the development of the
standard; and
    (4) The ANC literally embodies the concept of acidification as
posed by the basic principles of acid base chemistry and the
measurement method used to estimate ANC and, therefore, serves as a
direct index to protect against acidification.
    Ecological indicators of acidification in aquatic ecosystems can be
chemical or biological components of the ecosystem that are altered by
the acidifying effects of nitrogen and sulfur deposition. A desirable
ecological indicator for aquatic acidification is one that is
measurable or estimable, linked causally to deposition of nitrogen and
sulfur, and linked causally, either directly or indirectly to
ecological effects known or anticipated to adversely affect public
welfare.
    As summarized in chapter 2 of the PA, atmospheric deposition of
NOy and SOX causes aquatic acidification through
the input of strong acid anions (e.g., NO3- and
SO42-) that ultimately shifts the water chemistry
equilibrium toward increased hydrogen ion levels (or decreased pH). The
anions are deposited either directly to the aquatic ecosystem or
indirectly via transformation through soil nitrification processes and
subsequent drainage from terrestrial ecosystems. In other words, when
these anions are mobilized in the terrestrial soil, they can leach into
adjacent water bodies. Aquatic acidification is indicated by changes in
the surface water chemistry of ecosystems. In turn, the alteration of
surface water chemistry has been linked to negative effects on the
biotic integrity of freshwater ecosystems. There is a suite of chemical
indicators that could be used to assess the effects of acidifying
deposition on lake or stream acid-base chemistry. These indicators
include ANC; alkalinity (ALK); base neutralizing capacity, commonly
referred to as acidity (ACY); surface water pH; concentrations of
trivalent aluminum, Al\+3\; and concentrations of major anions
(SO42-, NO3-), cations
(Ca\2+\, Mg\+2\, K\+\), or sums of cations or anions.
    The ANC and ALK are very similar quantities and are used
interchangeably in the literature and for some of the analyses
presented in this document. Both ANC and ALK are defined as the amount
of strong acid required to reach a specified equivalence point. For
acid-base solutions, an equivalence point can be thought of as the
point at which the addition of strong acids (i.e., titration) is no
longer neutralized by the solution. This explains the term acid
neutralizing capacity, or ANC, as ANC relates directly to the capacity
of a system to neutralize acids. The differences between ANC and ALK
are based on operational definitions and subject to various
interpretations. The ANC is preferred over ALK as the body of
scientific evidence has focused on ANC and effects relationships. The
ALK is more widely associated with more general characterizations of
water quality such as the relative hardness of water associated with
carbonates.
    Indictors such as the concentrations of specific anions, cations,
or their groupings, while relevant to acidification processes, are not
robust acidification indicators as it is the relative balance of
cations and anions that is more directly associated with acidification.
That balance is captured by ANC and ALK. Acidity, ACY, is the converse
of ANC and indicates how much strong base it takes to reach an
equivalence point. Because ACY is not used in most ecosystem
assessments, the body of information relating ACY to effects is too
limited to serve as a basis for an appropriate ecological indicator.
Aluminum and other metals are causative toxic agents that directly
impair biological functions. However, Al, or metals in general, have
high variability in concentrations that can be linked to effects, often
at extremely low levels which in some cases approach detectability
limits, exhibit rapid transient responses, and are often confounded by
the presence of other toxic metals. These concerns limit the use of
metals as reliable and measurable ecological indicators. Hydrogen ion
(H\+\) concentrations, using their negative logarithmic values, or pH,
are well correlated with adverse effects, as discussed above in section
II.A, and determine the solubility of metals such as aluminum. However,
pH is not a preferred acidification indicator due to its highly
transient nature and other concerns, as discussed below.
    Having reasoned that ANC is a preferred indicator to ALK, ACY,
individual metals or groupings of ions, the PA considers the relative
merits of ANC compared to pH, which is a well recognized indicator of
acidity and a more direct causative agent with regard to adverse
effects. First, the linkage between ANC and pH is considered in
recognition of the causative association between pH and effects.
    The ANC is not the direct causative toxic agent impacting aquatic
species diversity. The scientific literature generally emphasizes the
links between pH and adverse effects as described above in section
II.A. It is important, therefore, to consider the extent to which ANC
and pH are well related from a mechanistic perspective as well as
through empirical evidence. The ANC and pH are co-dependent on each
other based on the requirement that all solutions are electrically
neutral, meaning that any solution must satisfy the condition that all
negatively charged species must be balanced by all positively charged
species. The ANC is defined as the difference between strong anions and
cations (US EPA, 2011, equation 7-13).
    While the chemistry can be complex, the co-dependency between ANC
and pH is explained by recognizing that positively charged hydrogen,
H\+\, is incorporated in the charge balance relationships related to
the overall solution chemistry which also defines ANC. The positive,
directional co-dependency (i.e., ANC and pH increase together) is
further explained in concept as ANC reflects how much strong acid
(i.e., how much hydrogen ion) it takes to titrate to an equivalence
point. Strong observed correlations between pH and ANC as described in
the PA support these mechanistic relationships.
    As discussed above in section II.A, there are well established
examples of ANC correlating strongly with a variety of ecological
effects which are summarized in the PA (US EPA, 2011, Table 3-1).
Because pH and ANC are well correlated and linearly dependent over the
pH ranges (4.5-6) where adverse ecological effects are observed,
evidence of clear associations exist between ANC and adverse ecological
effects as described in the PA. In large measure, this dependence
between pH

and ANC and the relationship of both pH and ANC to effects, speak
directly to the appropriateness of ANC with respect to its use as an
ecological indicator.
    Thus, there is a clear association between ANC and ecological
effects, although there is a more direct causal relationship between pH
and ecological effects. Nonetheless, ANC is preferred as an ecological
indicator based on its superior ability to provide a linkage with
deposition in a meaningful and quantifiable manner, a role that is
served far more effectively by ANC than by pH. While both ANC and pH
are clearly associated with the effects of concern, ANC is superior in
linking these effects to deposition.
    The PA notes that the basis for this conclusion is that acidifying
atmospheric deposition of nitrogen and sulfur is a direct input of
potential acidity (ACY), or, in terms of ANC, such deposition is
relevant to the major anions that reduce the capacity of a water body
to neutralize acidity. Consequently, there is a well defined linear
relationship between potential acidifying deposition and ANC. This ANC-
deposition relationship facilitates the linkage between ecosystem
models that calculate an ecological indicator and the atmospheric
deposition of NOy and SOX. On the other hand,
there is no direct linear relationship between deposition and pH. While
acid inputs from deposition lower pH, the relationship can be extremely
nonlinear and there is no direct connection from a modeling or mass
balance perspective between the amount of deposition entering a system
and pH. The term ``mass balance'' underlies the basic formulation of
any physical modeling construct, for atmospheric or aquatic systems,
and refers to the accounting of the flow of mass into a system, the
transformation to other forms, and the loss due to flow out of a system
and other removal processes. The ANC is a conserved property. This
means that ANC in a water body can be accounted for by knowledge of how
much ANC initially exists, how much flows in and is deposited, and how
much flows out. In contrast, hydrogen ion concentration in the water,
the basis for pH, is not a conserved property as its concentration is
affected by several factors such as temperature, atmospheric pressure,
mixing conditions of a water body, and the levels of several other
chemical species in the system. The disadvantage of pH lacking
conservative properties is that there is a very complex connection
between changes in ambient air concentrations of NOy and
SOX and pH.
    The discussion of basic water chemistry of natural systems in
chapter 2 of the PA provides further details on why pH is not a
conserved quantity and is subject to rapid transient response behavior
that makes it difficult to use as a reliable and functional ecological
indicator. The observed pH-to-ANC relationship (US EPA, 2011, figure 7-
2) partially explains the concern with pH responding too abruptly. In
the region where pH ranges roughly from 4.5 to 6 and is of greatest
relevance to effects (US EPA, 2011, figure 7-4), there clearly is more
sensitivity of pH to changes in ANC in the ANC range from approximately
0 to 50 [micro]eq/L. A focus on this part of the ANC-to-pH relationship
shows that ANC associates well with pH in a fairly linear manner.
However, the pH range from 4.5 to 6 also includes one of the very
steepest parts of the slope relating pH as a function of ANC, where ANC
ranges down below 0 [micro]eq/L, which is subject to very rapid change
in ANC, or deposition inputs. This part of the relationship coincides
with reduced levels of ANC and hence with reduced ability to neutralize
acids and moderate pH fluctuations. This response behavior can be
extended to considering how pH would change in response to deposition,
or ambient concentrations, of NOy and SOX, which
can be viewed as ``ANC-like'' inputs.
    In summary, because ANC clearly links both to biological effects of
aquatic acidification as well as to acidifying inputs of NOy
and SOX deposition, the PA concludes that ANC is an
appropriate ecological indicator for relating adverse aquatic ecosystem
effects to acidifying atmospheric deposition of SOx and
NOy, and is preferred to other potential indicators. In
reaching this conclusion, the PA notes that in its review of the first
draft PA, CASAC concluded that ``information on levels of ANC
protective to fish and other aquatic biota has been well developed and
presents probably the lowest level of uncertainty in the entire
methodology'' (Russell and Samet, 2010a). In its more recent review of
the second draft PA, CASAC agreed ``that acid neutralizing capacity is
an appropriate ecological measure for reflecting the effects of aquatic
acidification'' (Russell and Samet, 2010b; p. 4).
2. Linking ANC to Deposition
    There is evidence to support a quantified relationship between
deposition of nitrogen and sulfur and ANC. This relationship was
analyzed in the REA for two case study areas, the Adirondack and
Shenandoah Mountains, based on time-series modeling and observed
trends. In the REA analysis, long-term trends in surface water nitrate,
sulfate and ANC were modeled using MAGIC for the two case study areas.
These data were used to compare recent surface water conditions in 2006
with preindustrial conditions (i.e. preacidification 1860). The results
showed a marked increase in the number of acid impacted lakes,
characterized as a decrease in ANC levels, since the onset of
anthropogenic nitrogen and sulfur deposition, as discussed in chapter 2
of the PA.
    In the REA, more recent trends in ANC, over the period from 1990 to
2006, were assessed using monitoring data collected at the two case
study areas. In both case study areas, nitrate and sulfate deposition
decreased over this time period. In the Adirondack Mountains, this
corresponded to a decreased concentration of nitrate and sulfate in the
surface waters and an increase in ANC (U.S. EPA, 2009, section
<IP_ADDRESS>). In the Shenandoah Mountains, there was a slight decrease in
nitrate and sulfate concentration in surface waters corresponding to
modest increase in ANC from 50 [mu]eq/L in 1990 to 67 [mu]eq/L in 2006
(U.S. EPA, 2009, section <IP_ADDRESS>, Appendix 4, and section 3.4).
    In the REA, the quantified relationship between deposition and ANC
was investigated using ecosystem acidification models, also referred to
as acid balance models or critical loads models (U.S. EPA, 2011,
section 2 and U.S. EPA, 2009, section 4 and Appendix 4). These models
quantify the relationship between deposition of nitrogen and sulfur and
the resulting ANC in surface waters based on an ecosystem's inherent
generation of ANC and ability to neutralize nitrogen deposition through
biological and physical processes. A critical load is defined as the
amount of acidifying atmospheric deposition of nitrogen and sulfur
beyond which a target ANC is not reached. Relatively high critical load
values imply that an ecosystem can accommodate greater deposition
levels than lower critical loads for a specific target ANC level.
Ecosystem models that calculate critical loads form the basis for
linking deposition to ANC.
    As discussed in chapter 2 of the PA, both dynamic and steady state
models calculate ANC as a function of ecosystem attributes and
atmospheric nitrogen and sulfur deposition, and can be used to
calculate critical loads. Steady state models are time invariant and
reflect the long term consequences associated with an ecosystem
reaching equilibrium under a constant level of atmospheric deposition.
Dynamic models are time variant and take into account the time
dependencies inherent in ecosystem hydrology, soil and

biological processes. Dynamic models like MAGIC can provide the time
series response of ANC to deposition whereas steady state models
provide a single ANC relationship to any fixed deposition level.
Dynamic models naturally are more complex than steady state models as
they attempt to capture as much of the fundamental biogeochemical
processes as practicable, whereas steady state models depend on far
greater parameterization and generalization of processes that is
afforded, somewhat, by not having to accounting for temporal
variability.
    The PA notes that steady state models are capable of addressing the
question of what does it take to reach and sustain a specific level of
ANC. Dynamic models are also capable of addressing that question, but
can also address the question of how long it takes to achieve that
result. Dynamic models afford the ability for more comprehensive
treatment of a variety of processes throughout the surface, soil and
bedrock layers within an ecosystem. For example, steady state models
treat sulfate as a mobile anion throughout the system, meaning that the
sulfate that is deposited to a watershed enters the water column and is
not influenced by soil adsorption or cation exchange. Dynamic models
can incorporate these time variant processes. The use of a steady state
model treating sulfate as totally mobile does not necessarily conflict
with the possibility of sulfate acting as a less than mobile ion at
certain times. The steady state assumption is premised on the long term
behavior of sulfate which can undergo periods of net adsorption
followed by periods of net desorption which can balance out over time.
The PA recognizes that as the richness of the available data increases,
in terms of parameters and spatial resolution, the incorporation of
dynamic modeling approaches in the standard setting process should
become more feasible. In determining an appropriate modeling approach
for the development of a NAAQS in this review, the PA considers both
the relevance of the question addressed as well as the ability to
perform modeling that provides relevant information for geographic
areas across the country.
    Dynamic models require a large amount of catchment level-specific
data relative to steady state models. Because of the time invariant
nature of steady state models, the data requirements that integrate
across a broad spectrum of ecosystem processes is achievable and
available now at the national level. Water quality data to support
steady state models currently exist for developing a national data base
for modeling nearly 10,000 catchments in the contiguous U.S. In
contrast, the data needs to support dynamic models for national-scale
analyses simply are not available at this time. Further, the
information provided by steady state modeling would be sufficient to
develop and analyze alternative NAAQS and the kind of protection they
would afford. While it would be of interest to also obtain information
about how much time it would take for a target ANC level to be
achieved, the absence of such information does not preclude developing
and evaluating alternative NAAQS using the AAI structure. Based on the
above considerations, the PA concludes that at this time steady state
critical load modeling is an appropriate tool for linking long-term ANC
levels to atmospheric deposition of nitrogen and sulfur for development
of an AAI that has national applicability.
    A steady state model is used to define the critical load, which is
the amount of atmospheric deposition of nitrogen (N) and sulfur (S)
beyond which a target ANC is not achieved and sustained.\7\ It is
expressed as:
---------------------------------------------------------------------------

    \7\ This section discusses the linkages between deposition of
nitrogen and sulfur and ANC. Section III.B.3 then discusses the
linkages between atmospheric concentrations of NOY and
SOX and deposition of nitrogen and sulfur.

[GRAPHIC] [TIFF OMITTED] TP01AU11.024

---------------------------------------------------------------------------
Where:

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

    Equation III-1 is a modified expression that adopts the basic
formulation of the steady state models that are described in chapter 2
of the PA. More detailed discussion of the rationale, assumptions and
derivation of equation III-1, as well as all of the equations in this
section, are included in Appendix B of the PA. The equation simply
reflects the amount of deposition of nitrogen and sulfur from the
atmosphere, CLANClim(N + S), that is associated with a
sustainable long-term ANC target, [ANClim], given the
capacity of the natural system to generate ANC,
[BC]0-*, and the capacity of the natural system
to neutralize nitrogen deposition, Neco. This expression of critical
load is valid when nitrogen deposition is greater than Neco.\8\ The
runoff rate, Q, allows for balancing mass in the two environmental
mediums--atmosphere and catchment. This critical load expression can be
focused on a single water system or more broadly. To extend
applicability of the critical load expression (equation III-1) from the
catchment level to broader spatial areas, the terms Qr and
CLr, are used, which are the runoff rate and critical load,
respectively, of the region over which all the atmospheric terms in the
equation are defined.
---------------------------------------------------------------------------

    \8\ Because Neco is only relevant to nitrogen deposition, in
rare cases where Neco is greater than the total nitrogen deposition,
the critical load would be defined only in terms of acidifying
deposition of sulfur and the Neco term in equation III-1 would be
set to zero.
---------------------------------------------------------------------------

    In considering the contributions of SOx or
NOy species to acidification, it is useful to think of every
depositing nitrogen atom as supplying one equivalent charge unit and
every sulfur atom as depositing two charge units. The PA uses
equivalent charge per volume as a normalizing tool in place of the more
familiar metrics such as mass or moles per volume. This allows for a
clearer explanation of many of the relationships between atmospheric
and ecosystem processes that incorporate mass and volume unit
conventions somewhat specific to the environmental media of concern
(e.g., m\3\ for air and liter for liquid water). Equivalent charge
reflects the chemistry equilibrium fundamentals that assume
electroneutrality, or balancing charge where the sum of cations always
equals the sum of anions.
    As presented above, the terms S and N in the CLANClim (N
+ S) term broadly represent all species of sulfur or nitrogen that can
contribute to

acidifying deposition. This follows conventions used in the scientific
literature that addresses critical loads, and it reflects all possible
acidifying contributions from any sulfur or nitrogen species. For all
practical purposes, S reflects SOx as described above, the
sum of sulfur dioxide gas and particulate sulfate. However, N in
equation III-1 includes both oxidized forms, consistent with the
ambient indicator, NOy, in addition to the reduced nitrogen
species, ammonia and ammonium ion, referred to as NHx. The
NHX is included in the critical load formulation because it
contributes to potentially acidifying nitrogen deposition.
Consequently, from a mass balance or modeling perspective, the form of
the standard needs to account for NHX, as described below.
3. Linking Deposition to Ambient Air Indicators
    The last major component of the form illustrated in Figure III-1
addresses the linkage between deposition of nitrogen and sulfur and
concentrations of the ambient air indicators, NOY and
SOX. To link ambient air concentrations with deposition, the
PA defines a transference ratio, T, as the ratio of total wet and dry
deposition to ambient concentration, consistent with the area and time
period over which the standard is defined. To express deposition of
NOY and SOX in terms of NOY and
SOX ambient concentrations, two transference ratios were
defined, where TSOx equals the ratio of the combined dry and
wet deposition of SOx to the ambient air concentration of
SOx, and TNOY equals the ratio of the combined
dry and wet deposition of NOY to the ambient air
concentration of NOY.
    As described in chapter 7 of the PA, reduced forms of nitrogen
(NHx) are included in total nitrogen in the critical load
equation, III-1. Reduced forms of nitrogen are treated separately, as
are NOy and SOx, and the transference ratios are
applied. This results in the following critical load expression that is
defined explicitly in terms of the indicators NOY and
SOx:
[GRAPHIC] [TIFF OMITTED] TP01AU11.025

This is the same equation as III-1, with the deposition associated with
the critical load translated to deposition from ambient air
concentrations via transference ratios. In addition, deposition of
reduced nitrogen, oxidized nitrogen and oxidized sulfur are treated
separately.
    Transference ratios are a modeled construct, and therefore cannot
be compared directly to measurable quantities. There is an analogy to
deposition velocity, as a transference ratio is basically an aggregated
weighted average of the deposition velocities of all contributing
species across dry and wet deposition, and transference ratio units are
expressed as distance/time. However, wet deposition commonly is not
interpreted as the product of a concentration times a velocity. Direct
wet deposition observations are available which integrate all of the
processes, regardless of how well they may be understood, related to
wet deposition into a measurable quantity. There are reasonable
analogies between the processes governing dry and wet deposition, from
a fundamental mass transfer perspective. In both cases there is a
transfer of mass between the dry ambient phase and another medium,
either a surface or vegetation in the case of dry deposition, or a rain
droplet or cloud in the case of wet precipitation. The specific
thermodynamic properties and chemical/biological reactions that govern
the transfer of dry mass to plants or aqueous droplets differ, but
either process can be based on conceptualizing the product of a
concentration, or concentration difference, times a mass transfer
coefficient which is analogous to the basic dry deposition model: dry
deposition = concentration x velocity (U.S. EPA, 2011, Appendix F).
    Transference ratios require estimates of wet deposition of
NOy and SOX, dry deposition of NOY and
SOX, and ambient air concentrations of NOY and
SOx. Possible sources of information include model estimates
or a combination of model estimates and observations, recognizing that
dry deposition is a modeled quantity that can use observed or modeled
estimates of concentration. The limited amount of NOY
measurements in acid-sensitive areas as well as the combination of
representative NOY, SO2 and SO4
observations generally preclude the use of observations for development
of a standard that is applicable nationally.
    The PA considers a blending of observations and models to take
advantage of their relative strengths; e.g., combining the NADP wet
deposition observations, modeled dry deposition, and a mix of modeled
and observed concentrations, using the model for those species not
measured or measured with very sparse spatial coverage. A potential
disadvantage of mixing and matching observations and model estimates is
to lose consistency afforded by using just modeling alone. A modeling