difficult to measure and/or monetize the goods and services supplied by
ecosystems.  It can be informative in characterizing adversity to public
welfare to attempt to place an economic valuation on the set of goods
and services that have been identified with respect to a change in
policy however it must be noted that this valuation will be incomplete
and illustrative only.

	Knowledge about the relationships linking ambient concentrations and
ecosystem services is considered in the Policy Assessment as one method
by which to inform a policy judgment on a known or anticipated adverse
public welfare effect.  For example, a change in an ecosystem structure
and process, such as foliar injury, would be classified as an ecological
effect, with the associated changes in ecosystem services, such as
primary productivity, food availability, forest products, and aesthetics
(e.g., scenic viewing), classified as public welfare effects.
Additionally, changes in biodiversity would be classified as an
ecological effect, and the associated changes in ecosystem
services—productivity, existence (nonuse) value, recreational viewing
and aesthetics—would also be classified as public welfare effects.

	As described in Chapters 4 and 5 of the REA, case study analyses were
performed that link deposition in sensitive ecosystems to changes in a
given ecological indicator (e.g., for aquatic acidification, to changes
in acid neutralizing capacity [ANC]) and then to changes in ecosystems.
Appendix 8 of the REA links the changes in ecosystems to the services
they provide (e.g., fish species richness and its influence on
recreational fishing). To the extent possible for each targeted effect
area, the REA linked ambient concentrations of nitrogen and sulfur
(i.e., ambient air quality indicators) to deposition in sensitive
ecosystems (i.e., exposure pathways), and then to system response as
measured by a given ecological indicator (e.g., lake and stream
acidification as measured by ANC). The ecological effect (e.g., changes
in fish species richness) was then, where possible, associated with
changes in ecosystem services and the corresponding public welfare
effects (e.g., recreational fishing).

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.

Aquatic Acidification

	Acidification of aquatic ecosystems primarily affects the ecosystem
services that are derived from the fish and other aquatic life found in
surface waters.  In the northeastern United States, the surface waters
affected by acidification are not a major source of commercially raised
or caught fish; however, they are a source of food for some recreational
and subsistence fishers and for other consumers. Although data and
models are available for examining the effects on recreational fishing,
relatively little data are available for measuring the effects on
subsistence and other consumers.  Inland waters also provide aesthetic
and educational services along with non-use services, such as existence
value (protection and preservation with no expectation of direct use).
In general, inland surface waters such as lakes, rivers, and streams
also provide a number of regulating services, playing a role in
hydrological regimes and climate regulation. There is little evidence
that acidification of freshwaters in the northeastern United States has
significantly degraded these specific services; however, freshwater
ecosystems also provide biological control services by providing
environments that sustain delicate aquatic food chains.  The toxic
effects of acidification on fish and other aquatic life impair these
services by disrupting the trophic structure of surface waters.
Although it is difficult to quantify these services and how they are
affected by acidification, it is worth noting that some of these
services may be captured through measures of provisioning and cultural
services. For example, these biological control services may serve as
“intermediate” inputs that support the production of “final”
recreational fishing and other cultural services.

	As summarized in Chapter 4 of the Policy Assessment, recent studies
indicate that acidification of lakes and streams can result in
significant loss in economic value.  For example, data indicate that
more than 9% of adults in the northeastern part of the country
participate annually in freshwater fishing yielding 140 million
freshwater fishing days.  Each fishing day has an estimated average
value per day of $35. Therefore, the implied total annual value of
freshwater fishing in the northeastern United States was $5 billion in
2006.  Embedded in these numbers is a degree of harm to recreational
fishing services due to acidification that has occurred over time.
These harms have not been quantified on a regional scale; however, a
case study was conducted in the Adirondacks area (US EPA, 2011, section
4.4.2).

	In the Adirondacks case study, estimates of changes in recreational
fishing services were determined, as well as changes more broadly in
“cultural” ecosystem services (including recreational, aesthetic,
and nonuse services).  First, the MAGIC model (US EPA, 2009, Appendix 8
and section 2.2)  was applied to 44 lakes to predict what ANC levels
would be under both “business as usual” conditions (i.e., allowing
for some decline in deposition due to existing regulations) and
pre-emission (i.e., background) conditions.  Second, to estimate the
recreational fishing impacts of aquatic acidification in these lakes, an
existing model of recreational fishing demand and site choice was
applied.  This model predicts how recreational fishing patterns in the
Adirondacks would differ and how much higher the average annual value of
recreational fishing services would be for New York residents if lake
ANC levels corresponded to background (rather than business as usual)
conditions.  To estimate impacts on a broader category of cultural (and
some provisioning)  ecosystem services, results from the Banzhaf et al
(2006) valuation survey of New York residents were adapted and applied
to this context.  The survey used a contingent valuation approach to
estimate the average annual household WTP for future reductions in the
percent of Adirondack lakes impaired by acidification.  The focus of the
survey was on impacts on aquatic resources. Pretesting of the survey
indicated that respondents nonetheless tended to assume that benefits
would occur in the condition of birds and forests as well as in
recreational fishing.

	By extrapolating the 44 lake Adiriondack case study to all 3,000
Adirondack lakes and by applying the willingness to pay survey results
to all New York residents, the study estimated aggregated benefits
between $300 and $800 million annually for the equivalent of improving
lakes in the Adirondacks region to an ANC level of 50 µeq/L.  The REA
estimated 44% of the Adirondack lakes currently fall below an ANC of 50
µeq/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
µeq/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.

Terrestrial Acidification

	Chapter 4.4.3 and 4.4.4 of the Policy Assessment review several
economic studies of areas sensitive to terrestrial acidification.
Forests in the northeastern United States provide several important and
valuable provisioning ecosystem services, which are reflected in the
production and sales of tree products. Sugar maples are a particularly
important commercial hardwood tree species in the United States,
producing timber and maple syrup that provide hundreds of millions of
dollars in economic value annually.   Red spruce is also used in a
variety of wood products and provides up to $100 million in economic
value annually.  Although the data do not exist to directly link
acidification damages to economic values of lost recreational ecosystem
services in forests, these resources are valuable to the public.  A
recent study, reviewed in the Policy Assessment, suggests that the total
annual value of recreational off-road driving was more than $9 billion
and the value of hunting and wildlife viewing was more than $4 billion
each in the northeastern United States.   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.

Nutrient Enrichment

	Chapter 4.4.5 and 4.4.6 of the Policy Assessment summarize economic
studies of east coast estuaries affected by nutrient over-enrichment or
eutrophication.  Estuaries in the eastern United States are important
for fish and shellfish production. The estuaries are capable of
supporting large stocks of resident commercial species, and they serve
as the breeding grounds and interim habitat for several migratory
species. To provide an indication of the magnitude of provisioning
services associated with coastal fisheries, from 2005 to 2007, the
average value of total catch was $1.5 billion per year in 15 East Coast
states.  Estuaries also provide an important and substantial variety of
cultural ecosystem services, including water-based recreational and
aesthetic services. For example, data indicate that 4.8% of the
population in coastal states from North Carolina to Massachusetts
participated in saltwater fishing, with a total of 26 million saltwater
fishing days in 2006.  Based on estimates in the Policy Assessment,
total recreational value from these saltwater fishing days was
approximately $1.3 billion.  Recreational participation estimates for
1999–2000 showed almost 6 million individuals participated in
motorboating in coastal states from North Carolina to Massachusetts. The
aggregate value of these coastal motorboating outings was $2 billion per
year.  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.

	Terrestrial ecosystems can also suffer from nutrient over-enrichment.
Each ecosystem is different in its composition of species and nutrient
requirements.  Changes to individual ecosystems from changes in nitrogen
deposition can be hard to assess economically.  Relative recreational
values are often determined by public use information.  Chapter 4.4.7 of
the Policy Assessment reviewed studies related to park use in
California.  Data from California State Parks indicate that in 2002,
68.7% of adult residents participated in trail hiking for an average of
24.1 days per year. The analyses in the Policy Assessment indicate that
the aggregate annual benefit for California residents from trail hiking
in 2007 was $11.59 billion.  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.

	The Policy Assessment also identified fire regulation as a service that
could be affected by nutrient over-enrichment of the Coastal Sage Scrub
and Mixed Conifer Forest ecosystems by encouraging growth of more
flammable grasses, increasing fuel loads, and altering the fire cycle.
Over the 5-year period from 2004 to 2008, Southern California
experienced, on average, over 4,000 fires per year, burning, on average,
over 400,000 acres per year.  It is not possible at this time to
quantify the contribution of nitrogen deposition, among many other
factors, to increased fire risk.

Summary

	Adversity to public welfare can be understood not only by looking at
how deposition of oxides of nitrogen and sulfur affect the ecological
functions of an ecosystem (see II.A.), but also and then by
understanding the ecosystem services that are degraded.  The monetized
value of the ecosystem services provided by ecosystems that are
sensitive to deposition of oxides of nitrogen and sulfur are in the
billions of dollars each year, though it is not possible to quantify or
monetize at this time the effects on these values of acid 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.  Important biodiversity is 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 and nitrogen and sulfur standard is
whether, in view of the scientific evidence reflected in the ISA,
additional information on exposure and risk discussed in the REA, and
conclusions drawn from the Policy Assessment, the existing standards
provide adequate protection.   The Administrator therefore, has
considered the extent to which the current standards are adequate for
the protection of public welfare.    Having reached theis general
conclusion that aquatic and terrestrial ecosystems can be degrade by
deposition of oxides of nitrogen and sulfur, it is then necessary to
first evaluate the appropriateness of the current standards to address
the ecological effects of oxides of nitrogen and sulfur as well as the
adequacy of the current secondary standards for oxides of nitrogen and
sulfur to provide requisite protection by considering to what degree
risks to sensitive ecosystems would be expected to occur in areas that
meet the current standards.  Conclusions regarding the adequacy of the
current standards are based on the available ecological effects,
exposure and risk-based evidence.   In evaluating the strength of this
information, EPA has taken into account the uncertainties and
limitations in the scientific evidence.  This section addresses the
adequacy of the current standards to protect against direct exposure
effects on plants from oxides of nitrogen and sulfur, the
appropriateness of the current structure of the standards to address
deposition-related effects of oxides of nitrogen and sulfur on sensitive
ecosystems and finally, the adequacy of such standards to protect
against adverse effects related to the deposition of oxides of nitrogen
and sulfur.

1.	Adequacy of the Current Standards for Direct Effects

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

Appropriateness and Adequacy of the Current Standards for
Deposition-related Effects

	This section addresses two concepts necessary to evaluate the current
standards on 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
relates evaluates the current standards and in terms of adequacy of
protection.

Appropriateness

	The ISA has established that the major effects of concern for this
review of the oxides of nitrogen and sulfur standards are associated
with deposition of N and S caused by atmospheric concentrations of
oxides of nitrogen and sulfur.  The current standards are not directed
toward depositional effects, and none of the elements of the current
NAAQS – indicator, form, averaging time, and level – are suited for
addressing the effects of N and S deposition.

 issues arise that call into question the ecological relevance of the
structure of the current secondary standards for oxides of nitrogen and
sulfur.

 standards do not utilize appropriate atmospheric indicators.  NO2 and
SO2 are used as the component of oxides of nitrogen and sulfur that are
measured, but they do not provide a complete link to the direct effects
on ecosystems from deposition of oxides of nitrogen and sulfur as they
do not capture all relevant chemical species of oxidized nitrogen and
oxidized sulfur that contribute to deposition.  The ISA provides
evidence that deposition related effects are linked with total nitrogen
and total sulfur deposition, and thus all forms of oxidized nitrogen and
oxidized sulfur that are deposited will contribute to effects on
ecosystems.  Thus, by using atmospheric NO2 and SO2 concentrations as
indicators, the current standards address only a fraction of total
atmospheric oxides of nitrogen and sulfur, and do not take into account
the effects from deposition of total atmospheric oxides of nitrogen and
sulfur.  This suggests that more comprehensive atmospheric indicators
should be considered in designing ecologically relevant standards.

Current standards reflect separate assessments of the two individual
pollutants, NO2 and SO2, rather than assessing the joint impacts of
deposition to ecosystems.  Recognizing the role that each pollutant
plays in jointly affecting ecosystem indicators, functions, and services
is vital to developing a meaningful standard.  The clearest example of
this interaction is in assessment of the impacts of acidifying
deposition on aquatic ecosystems.  Acidification in an aquatic ecosystem
depends on the total acidifying potential of the deposition of both N
and S from both atmospheric deposition of oxides of nitrogen and sulfur
as well as the inputs from other sources of N and S such as reduced
nitrogen and non-atmospheric sources. It is the joint impact of the two
pollutants that determines the ultimate effect on organisms within the
ecosystem, and critical ecosystem functions such as habitat provision
and biodiversity.  Standards that are set independently are less able to
account for the contribution of the other pollutant.  This suggests that
interactions between oxides of nitrogen and oxides of sulfur should be a
critical element of the conceptual framework for ecologically relevant
standards.  There are also important interactions between oxides of
nitrogen and sulfur and reduced forms of nitrogen, which also contribute
to acidification and nutrient enrichment.  Although the standards do not
directly address reduced forms of nitrogen in the atmosphere, e.g. they
do not require specific levels of reduced nitrogen, iIt is important
that the structure of the standards address the role of reduced nitrogen
in determining the ecological effects resulting from deposition of
atmospheric oxides of nitrogen and sulfur.  Consideration will also have
to be given to total loadings as ecosystems respond to all sources of N
and S.

Based on the discussioned summarized above, the Policy Assessment
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.

Adequacy of Protection

 In addition, these levels based on conclusions in the REA will not
decline in the future to levels below which it is reasonable to
anticipate effects.

	 In determining the adequacy of the current secondary standards for
oxides of nitrogen and sulfur the Policy Assessment considered the
extent to which ambient deposition contributes to loadings in
ecosystems.  Since the last review of the secondary standard for oxides
of nitrogen, a great deal of information on the contribution of
atmospheric deposition associated with ambient oxides of nitrogen has
become available.  The REA presents a thorough assessment of the
contribution of oxidized nitrogen to nitrogen deposition throughout the
U.S., and the relative contributions of ambient oxidized and reduced
forms of nitrogen.  The REA concludes that based on that analysis,
ambient oxides of nitrogen are a significant component of atmospheric
nitrogen deposition, even in areas with relatively high rates of
deposition of reduced nitrogen.  In addition, atmospheric deposition of
oxidized nitrogen contributes significantly to total nitrogen loadings
in nitrogen sensitive ecosystems.

, the ISA indicates that atmospheric N deposition is the main source of
new anthropogenic N to most headwater streams, high elevation lakes, and
low-order streams. Atmospheric N deposition contributes to the total N
load in terrestrial, wetland, freshwater, and estuarine ecosystems that
receive N through multiple pathways.  In several large estuarine
systems, including the Chesapeake Bay, atmospheric deposition accounts
for between 10 and 40 percent of total nitrogen loadings (US EPA, 2008).

	Atmospheric concentrations of oxides of sulfur account for nearly all S
deposition in the US.  For the period 2004–2006, mean S deposition in
the U.S. was greatest east of the Mississippi River with the highest
deposition amount, 21.3 kg S/ha-yr, in the Ohio River Valley where most
recording stations reported 3 year averages >10 kg S/ha-yr. Numerous
other stations in the East reported S deposition >5 kg S/ha-yr. Total S
deposition in the U.S. west of the 100th meridian was relatively low,
with all recording stations reporting <2 kg S/ha-yr and many reporting
<1 kg S/ha-yr. S was primarily deposited in the form of wet SO4 2−
followed in decreasing order by a smaller proportion of dry SO2 and a
much smaller proportion of deposition as dry SO42−.

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

Aquatic acidification

	The focus of the REA case studies was on determining whether deposition
of sulfur and oxidized nitrogen in locations where ambient oxides of
nitrogen and sulfur were at or below the current standards was resulting
in acidification and related effects, including episodic acidification
and mercury methylation.  Based on the case studies conducted for lakes
in the Adirondacks and streams in Shenandoah National Park (case studies
are discussed more fully in section IIB and US EPA, 2009), there is
significant risk to acid sensitive aquatic ecosystems at atmospheric
concentrations of oxides of nitrogen and sulfur at or below the current
standards.  The REA also supports strongly a relationship between
atmospheric deposition of oxides of nitrogen and sulfur and loss of ANC
in sensitive ecosystems and 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.

 This information indicates that almost half of the 44 lakes in the
Adirondacks case study area are at an elevated concern levels, and
almost a third are at a severe concern level. These levels are
associated with greatly diminished fish species diversity, and losses in
the health and reproductive capacity of remaining populations. Based on
assessments of the relationship between number of fish species and ANC
level in both the Adirondacks and Shenandoah areas, the number of fish
species is decreased by over half at an ANC level of 20 μeq/L relative
to an ANC level at 100 μeq/L (US EPA, 2009, Figure 4.2-1).  When
extrapolated to the full population of lakes in the Adirondacks area
using weights based on the EMAP probability survey (US EPA, 2009,
section <IP_ADDRESS>), 36 percent of lakes exceeded the critical load for an
ANC of 50 μeq/L and 13 percent of lakes exceeded the critical load for
an ANC of 20 μeq/L.

As with the Adirondacks area, this information suggests that significant
numbers of sensitive streams in the Shenandoah area are at risk of
adverse impacts on fish populations under recent conditions. Many other
streams in the Shenandoah area are 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 Policy Assessment all conclude that the current
standards are not adequate to protect against the adverse impacts of
aquatic acidification on sensitive ecosystems.  In the ISA it is noted
that significant portions of the U.S. are acid sensitive, and that
current deposition levels exceed those that would allow recovery of the
most acid sensitive lakes in the Adirondacks (US EPA, 2008, Executive
Summary). In addition, because of past loadings, areas of the Shenandoah
are sensitive to current deposition levels (US EPA, 2008, Executive
Summary). A recent survey, as reported in the ISA, found sensitive
streams in many locations in the U.S., including the Appalachian
Mountains, the Coastal Plain, and the Mountainous West (US EPA, 2008,
section <IP_ADDRESS>). In these sensitive areas, between 1 and 6 percent of
stream kilometers are chronically acidified. The REA further concludes
that both the Adirondack and Shenandoah case study areas are currently
receiving deposition from ambient oxides of nitrogen and sulfur in
excess of their ability to neutralize such inputs.  In addition based on
the current emission scenarios, forecast modeling out to the year 2020
as well as 2050 indicates a large number of streams in these areas will
still be adversely impacted (section IIB).  Based on these
considerations, the Policy Assessment concludes that the current
secondary NAAQS for oxides of nitrogen and sulfur do not provide
adequate protection of sensitive ecosystems with regard to aquatic
acidification.

Terrestrial acidification

	Based on the terrestrial acidification case studies (Kane Experimental
Forest in Pennsylvania and Hubbard Brook Experimental Forest (case study
locations described in section IIB) on sugar maple and red spruce
habitat, the REA concludes that there is significant risk to sensitive
terrestrial ecosystems from acidification at atmospheric concentrations
of NOx and SOX at or below the current standards.    The ecological
indicator selected for terrestrial acidification is the base cation to
aluminum ratio (BC:Al), which has been linked to tree health and growth.
 The results of the REA strongly support a relationship between
atmospheric deposition of oxides of nitrogen and sulfur and BC:Al, and
that BC:Al is a good indicator of terrestrial acidification.  At levels
of deposition associated with oxides of nitrogen and sulfur
concentrations at or below the current standards, BC:Al levels are
expected to be below benchmark values that are associated with
significant effects on  tree health and growth. Such degradation of
terrestrial ecosystems could affect ecosystem services such as habitat
provisioning, endangered species, goods production (timber, syrup, etc.)
and many 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>).

.  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.

Terrestrial nutrient enrichment

	Nutrient enrichment effects are due to nitrogen loadings from both
atmospheric and non-atmospheric sources. Evaluation of nutrient
enrichment effects requires an understanding that nutrient inputs are
essential to ecosystem health and that specific long term levels of
nutrients in a system affect the types of species that occur over long
periods of time.  Short term additions of nutrients can affect species
competition, and even small additions of nitrogen in areas that are
traditionally nutrient poor can have significant impacts on productivity
as well as species composition.    Most ecosystems in the United States
are nitrogen-limited, so regional decreases in emissions and deposition
of airborne nitrogen compounds could lead to some decrease in growth of
the vegetation that surrounds the targeted aquatic system but as
discussed below evidence for this is mixed. Whether these changes in
plant growth are seen as beneficial or adverse will depend on the nature
of the ecosystem being assessed.

	Information on the effects of changes in nitrogen deposition on
forestlands and other terrestrial ecosystems is very limited. The
multiplicity of factors affecting forests, including other potential
stressors such as ozone, and limiting factors such as moisture and other
nutrients, confound assessments of marginal changes in any one stressor
or nutrient in forest ecosystems.  The ISA notes that only a fraction of
the deposited nitrogen is taken up by the forests, most of the nitrogen
is retained in the soils (US EPA, 2008, section <IP_ADDRESS>). In addition,
the ISA indicates that forest management practices can significantly
affect the nitrogen cycling within a forest ecosystem, and as such, the
response of managed forests to NOx deposition will be variable depending
on the forest management practices employed in a given forest ecosystem
(US EPA, 2008, Annex C C.6.3).  Increases in the availability of
nitrogen in N-limited forests via atmospheric deposition could increase
forest production over large non-managed areas, but the evidence is
mixed, with some studies showing increased production and other showing
little effect on wood production (US EPA, 2008, section 3.3.9). Because
leaching of nitrate can promote cation losses, which in some cases
create nutrient imbalances, slower growth and lessened disease and
freezing tolerances for forest trees, the net effect of increased N on
forests in the U.S. is uncertain (US EPA, 2008, section 3.3.9).

, the community of lichens begins to change from acidophytic to tolerant
species; at 5.2 kg N/ha-yr, the typical dominance by acidophytic species
no longer occurs; and at 10.2 kg N/ha-yr, acidophytic lichens are
totally lost from the community. Additional studies in the Colorado
Front Range of the Rocky Mountain National Park support these findings.
These three values (3.1, 5.2, and 10.2 kg/ha-yr) are one set of
ecologically meaningful benchmarks for the mixed conifer forest (MCF) of
the pacific coast regions. Nearly all of the known sensitive communities
receive total nitrogen deposition levels above the 3.1 N kg/ha-yr
ecological benchmark according to the12 km, 2002 CMAQ/NADP data, with
the exception of the easternmost Sierra Nevadas. MCFs in the southern
portion of the Sierra Nevada forests and nearly all MCF communities in
the San Bernardino forests receive total nitrogen deposition levels
above the 5.2 N kg/ha-yr ecological benchmark.

Aquatic nutrient enrichment

.  In addition, this type of indicator does not reflect the impact of
nitrogen deposition in conjunction with other sources of nitrogen.

	Based on the above considerations, the REA concludes that the ASSETS EI
is not an appropriate ecological indicator for estuarine aquatic
eutrophication and that additional analysis is required to develop an
appropriate indicator for determining the appropriate levels of
protection from N nutrient enrichment effects in estuaries related to
deposition of oxides of nitrogen.  As a result, EPA is unable to make a
determination as to the adequacy of the existing secondary oxides of
nitrogen standard in protecting public welfare from N nutrient
enrichment effects in estuarine aquatic ecosystems.

Other effects

neurotoxic contaminant.  The production of meaningful amounts of
methylmercury (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 methylmercury requires the presence of sulfate and
mercury, but the amount of methylmercury produced varies with oxygen
content, temperature, pH, and supply of labile organic carbon (US EPA,
2008, section 3.4). In watersheds where changes in sulfate deposition
did not produce an effect, one or several of those interacting factors
were not in the range required for meaningful methylation to occur (US
EPA, 2008, section 3.4). Watersheds with conditions known to be
conducive to mercury methylation can be found in the northeastern United
States and southeastern Canada (US EPA, 2009, section 6).

	With respect to sulfur deposition and mercury methylation, the final
ISA determined: The evidence is sufficient to infer a causal
relationship between sulfur deposition and increased mercury methylation
in wetlands and aquatic environments. However, EPA did not conduct a
quantitative assessment of the risks associated with increased mercury
methylation under current conditions. As such, EPA is unable to make a
determination as to the adequacy of the existing SO2 secondary standards
in protecting against welfare effects associated with increased mercury
methylation.

vi. Summary of Adequacy Considerations

	In summary, the Policy Assessment concludes that currently available
scientific evidence and assessments clearly call into question the
adequacy of the current standards with regard to deposition-related
effects on sensitive aquatic and terrestrial ecosystems, including
acidification and nutrient enrichment.  Further, the Policy Assessment
recognizes that the elements of the current standards -- indicator,
averaging time, level and form – are not ecologically relevant, and
are thus not appropriate for standards designed to provide such
protection.  Thus, the Policy Assessment concludes that consideration
should be given to establishing a new ecologically relevant
multi-pollutant, multimedia standard to provide appropriate protection
from deposition-related ecological effects of oxides of nitrogen and
sulfur on sensitive ecosystems, with a focus on protecting against
adverse effects associated with acidifying deposition in sensitive
aquatic ecosystems.

CASAC Views

	In a letter to the Administrator (Russell and Samet 2011a), the CASAC
Oxides of Nitrogen and Oxides of Sulfur Panel, with full endorsement of
the chartered CASAC, unanimously concluded that:

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

	 In addition, with regard to the joint consideration of both oxides of
nitrogen and oxides of sulfur as well as the consideration of deposition
related effects, CASAC concluded that the Policy Assessment had
developed a credible methodology for considering such effects.  The
Panel stated that “the Policy Assessment develops a framework for a
multi-pollutant, multimedia standard that is ecologically relevant and
reflects the combined impacts of these two pollutants as they deposit to
sensitive aquatic ecosystems.”

Administrator’s Proposed Conclusions Concerning Adequacy of Current
Standard

	Based on the above considerations and taking into account CASAC advice,
the Administrator recognizes that the purpose of the secondary standard
is to protect against “adverse” effects resulting from exposure to
oxides of nitrogen and sulfur, discussed above in section IIA. The
Administrator also recognizes the need for conclusions both 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 Policy Assessment, and the
views of CASAC and members of the public.  On these bases, the
Administrator concludes that the current secondary standards are
adequate to protect against direct phytotoxic effects on vegetation.
Thus, the Administrator proposes to retain the current secondary
standard for oxides of nitrogen at 53 ppb, 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.

	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 Policy
Assessment which highlighted the sufficiency of the quantity and quality
of the available evidence and assessments associated with aquatic
acidification relative to the information and assessments available for
other deposition-related effects, including terrestrial acidification
and aquatic and terrestrial nutrient enrichment.  Based on its review of
these documents, CASAC agreed that aquatic acidification should be the
focus for developing a new multi-pollutant standard in this review.  In
reaching conclusions about an air quality standard designed to address
deposition-related aquatic acidification effects, the Administrator also
recognizes that such a standard may also provide some degree of
protection against other deposition-related effects.

	As discussed in chapter 7 of the Policy Assessment, the development of
a new 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 to public welfare, while recognizing the complex
and heterogeneous interactions between ambient air concentrations of
nitrogen and sulfur oxides, the related deposition of nitrogen and
sulfur, and associated ecological responses.  The development of such a
standard also needs to take into account the limitations and
uncertainties in the available information and analyses upon which
characterization of such interactions are based.  The approach used in
the Policy Assessment also recognizes that while such a standard would
be national in scope and coverage, the effects to public welfare from
aquatic acidification will not occur to the same extent in all locations
in the U.S., given the inherent variability of the responses of aquatic
systems to the effects of acidifying deposition.

	As discussed above in section II, many locations in the U.S. are
naturally protected against acid deposition due to underlying geological
conditions.  Likewise, some locations in the U.S., including lands
managed for commercial agriculture and forestry, are not likely to be
negatively impacted by current levels of nitrogen and sulfur deposition.
 As a result, while a new ecologically relevant secondary standard would
apply everywhere, it would be structured to account for differences in
the sensitivity of ecosystems across the country.  This would allow for
appropriate protection of sensitive aquatic ecosystems, which are
relatively pristine and wild and generally in rural areas, and the
services provided by such sensitive ecosystems, without requiring more
protection than is needed elsewhere.

	As discussed below, the multi-pollutant standard developed in the
Policy Assessment would employ (1) NOy and SOx as the atmospheric
ambient air indicators; (2) a form that takes into account variable
factors, such as atmospheric and ecosystem conditions that modify the
amounts of deposited nitrogen and sulfur; the distinction between
oxidized and reduced forms of nitrogen; effects of deposited nitrogen
and sulfur on aquatic ecosystems in terms of the ecological indicator
ANC; and the representativeness of water bodies within a defined spatial
area; (3) a multi-year averaging time, and (4) a standard level defined
in terms of a single, national target ANC value that, in the context of
the above form, identifies the levels of concentrations of NOy and SOx
in the ambient air that would meet the standard.  The form of such a
standard has been defined by an index, termed an aquatic acidification
index (AAI), which reflects the relationship between ambient
concentrations of NOy and SOx and aquatic acidification effects that
result from nitrogen and sulfur deposition related to these ambient
concentrations.

	In presenting the considerations associated with such an air quality
standard to address deposition-related aquatic acidification effects,
the following sections focus on each element of the standard, including
indicator (section III.A), form (section III.B), averaging time (section
III.C), and level (section III.D).   Alternative combinations of levels
and forms are discussed in section III.E.  Considerations related to
important uncertainties inherent in such an approach are discussed in
section III.F.  Advice from CASAC on such a new standard is presented in
section III.G.  The Administrator’s proposed decisions on such a new
standard are presented in section III.H.

A.	Ambient Air Indicators

	In considering alternative ambient air indicators, the Policy
Assessment primarily focuses on the important attribute of association.
Association in a broad sense refers to how well an ambient air indicator
relates to the ecological effects of interest by virtue of both the
framework that links the ambient indicator and effects and the empirical
evidence that quantifies the linkages.  The Policy Assessment also
considers how measurable or quantifiable an indicator is to enable its
use as an effective indicator of relevant ambient air concentrations.

This includes both the species of oxides of nitrogen and sulfur that are
directly emitted as well as species transformed in the atmosphere from
oxides of nitrogen and sulfur that retain the nitrogen and sulfur atoms
from directly emitted oxides of nitrogen and sulfur.   All of these
compounds are associated with oxides of nitrogen and sulfur in the
ambient air and can contribute to acidifying deposition.

	The Policy Assessment focuses in particular on the various compounds
with nitrogen or sulfur atoms that are associated with oxides of
nitrogen and sulfur, because the acidifying potential is specific to
nitrogen and sulfur, and not other atoms (e.g., H, C, O) whether derived
from the original source of oxides of nitrogen and sulfur emissions or
from atmospheric transformations.  For example, the acidifying potential
of each molecule of NO2, NO, HNO3 or PAN is identical, as is the
potential for each molecule of SO2 or ion of particulate sulfate, p-SO4.
  Each atom of sulfur affords twice the acidifying potential of each
atom of nitrogen.

1.	Oxides of Sulfur

 does represent virtually the entire ambient air mass of sulfur that
contributes to acidification.  In addition to accounting for virtually
all the potential for acidification from oxidized sulfur in the ambient
air, there are reliable methods to monitor the concentrations of SO2 and
particulate SO4.  In addition, much of the data used to develop the
technical basis for the standard developed in the Policy Assessment is
based on monitoring or modeling of these species.  The Policy Assessment
concludes that the sum of SO2 and SO4, referred to as SOx, are
appropriate ambient air indicators of oxides of sulfur because they
represent virtually all of the acidification potential of ambient air
oxides of sulfur and there are reliable methods suitable for measuring
SO2 and SO4.

2.	Oxides of Nitrogen

	As discussed in the Policy Assessment (US EPA, 2011,section 7.1.2),
total reactive oxidized nitrogen, NOy, as defined in chapter 2 of the
Policy Assessment, incorporates basically all of the oxidized nitrogen
species that have acidifying potential and as such, NOy should be
considered as an appropriate indicator for oxides of nitrogen.  NOy is
an aggregate measure of s NO and NO2 and all of the reactive oxidized
products of NO and NO2.  That is, NOy is a group of nitrogen compounds
in which all of the compounds are either an oxide of nitrogen or
compounds in which the nitrogen atoms came from oxides of nitrogen.  NOy
is especially relevant as an ambient indicator for acidification in that
it both relates to the oxides of nitrogen in the ambient air and also
represents the acidification potential of all oxidized nitrogen species
in the ambient air, whether an oxide of nitrogen or derived from oxides
of nitrogen.

	There are currently available reliable methods of measuring aggregate
NOy.  The term “aggregate” measure means that the NOY, as measured,
is not based on measuring each individual species of NOy and calculating
an NOy value by summing the individual species.   Rather, as described
in chapter 2 of the Policy Assessment, current measurement techniques
process all of the individual NOy species to produce a single aggregate
measure of all of the nitrogen atoms associated with any NOy species.
Consequently, the NOy measurement effectively provides the sum of all
individual species, but the identity of the individual species is lost.
As discussed above, the accounting for the individual nitrogen atoms is
an accounting of the ambient air acidification potential of oxides of
nitrogen and their transformation products and therefore the most
relevant ambient indicator for aquatic acidification effects associated
with oxides of nitrogen.

	This loss of the information on individual species motivated
consideration of alternative or more narrowly defined indicators for
oxides of nitrogen in the Policy Assessment.  Consideration of a subset
of NOy species was based on the following reasoning.  First, the actual
dry deposition of nitrogen is determined on an individual species basis
by multiplying the species concentration times a species-specific
deposition velocity and then summed to develop an estimate of total dry
deposition.  Consequently, the use of individual ambient species has the
potential to be more consistent with the underlying science of
deposition and, therefore, has the potential to allow for a more
rigorous evaluation of dry deposition with specialized field studies.
In addition, there has been a suggestion of focusing only on the most
quickly depositing NOy species, such as HNO3, as contributions from
other NOy species such as NO2 may be negligible.  These alternative
indicators are discussed below.

	The Policy Assessment considers the relative merits of using each
individual NOy species as part of a group of indicators.  In so doing,
it was first noted that dry deposition of NOy is treated as the sum of
the deposition of each individual species in advanced process-based air
quality models like CMAQ, as described in chapter 2 of the Policy
Assessment.  Conceptually one could extend this process-based approach
by using all NOY species individually as separate indicators for oxides
of nitrogen and requiring, for example, measurements of each of the
species, including the dominant species of HNO3, particulate nitrate
(p-NO3), true NO2, NO, and PAN.   The potential attraction of using
individual species would be the reliance on actual deposition
velocities.  This could have more physical meaning in comparison to a
constructed model of aggregate deposition of NOy, which is difficult to
evaluate with observations because of the assimilation of many species
with disparate deposition behavior.  The Policy Assessment notes that
the major drawback of using individual NOy species as the indicators is
the lack of reliable measurement techniques, especially for PAN and NO2
in rural locations, which renders the use of virtually any individual
NOy species, except for NO and perhaps p-NO3, as functionally inadequate
from a measurement perspective.

	The Policy Assessment next considered the relative merits of using a
subset of NOy species as the indicators for oxides of nitrogen, as was
discussed above for oxides of sulfur.  To the extent that certain
species provide relatively minor contributions to total NOy deposition,
it may be appropriate to consider excluding them as part of the
indicator.  As discussed in chapter 2 of the Policy Assessment, each
nitrogen species within the array of NOy species has species-specific
dry deposition velocities.   For example, the deposition velocity of
HNO3 is much greater than the velocity for nitrogen dioxide and,
consequently, for a similar ambient air concentration, HNO3 contributes
more deposition of acidifying nitrogen relative to nitrogen dioxide.  In
transitioning from source-oriented urban locations to rural
environments, the ratio of the concentrations of HNO3 and PAN to NO2
increases.

 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).

.

B.	Form

	Based on the evidence of the aquatic acidification effects caused by
the deposition of NOy and SOx, the Policy Assessment (US EPA,
2011,section 7.2) presents the development of a new form that is
ecologically relevant for addressing such effects.  The conceptual
design for the form of such a standard includes three main components:
an ecological indicator, deposition metrics that relate to the
ecological indicator, and a function that relates ambient air indicators
to deposition metrics.  Collectively, these three components link the
ecological indicator to ambient air indicators, as illustrated above in
Fig II-1.

	The simplified flow diagram in Figure II-1 compresses the various
atmospheric, biological, and geochemical processes associated with
acidifying deposition to aquatic ecosystems into a simplified conceptual
picture.  The ecological indicator (left box) is related to atmospheric
deposition through biogeochemical ecosystem models (middle box), which
associate a target deposition load to a target ecological indicator.
Once a target deposition is established, associated allowable air
concentrations are determined (right box) through the relationships
between concentration and deposition that are embodied in air quality
models such as CMAQ.   The following discussion describes the
development and rationale for each of these components, as well as the
integration of these components into the full expression of the form of
the standard using the concept of a national AAI that represents a
target ANC level as a function of ambient air concentrations.  Spatial
aggregation issues associated with defining each of the terms of this
index are also addressed below.

 	The AAI is designed to be an ecologically relevant form of the
standard that determines the levels of NOy and SOx in the ambient air
that would achieve a target ANC limit for the U.S.  The intent of the
AAI is to weight atmospheric concentrations of oxides of nitrogen and
sulfur by their propensity to contribute to acidification through
deposition, given the fundamental acidifying potential of each
pollutant, and to take into account the ecological factors that govern
acid sensitivity in different ecosystems.  The index also accounts for
the contribution of reduced nitrogen to acidification.  Thus, the AAI
encompasses those attributes of specific relevance to protecting
ecosystems from the acidifying potential of ambient air concentrations
of NOy and SOx.

1.	Ecological Indicator

 	In considering alternative ecological indicators, the Policy
Assessment again primarily focuses on the attribute of association.  In
the case of an ecological indicator for aquatic acidification,
association refers to the relationship between the indicator and adverse
effects as discussed in section II.  Because of the conceptual structure
of the form of this standard (Figure III-1), this particular ecological
indicator must also link up in a meaningful and quantifiable manner with
acidifying atmospheric deposition.  In effect, the ecological indicator
for aquatic acidification is the bridge between biological impairment
and deposition of NOy and SOx.

	This section presents the rationale in the Policy Assessment for
selecting acid neutralizing capacity (ANC) as the appropriate ecological
indicator for consideration.  Recognizing that ANC is not itself the
causative or toxic agent for adverse aquatic acidification effects, the
rationale for using ANC as the relevant ecological indicator is based on
the following:

ANC is directly associated with the causative agents, pH and dissolved
aluminum, both through empirical evidence and mechanistic relationships;

Empirical evidence shows very clear and strong relationships between
adverse effects and ANC;

ANC is a more reliable indicator from a modeling perspective, allowing
use of a body of studies and technical analyses related to ANC and
acidification to inform the development of the standard; and

ANC literally embodies the concept of acidification as posed by the
basic principles of acid base chemistry and the measurement method used
to estimate ANC and, therefore, serves as a direct index to protect
against acidification.

	Ecological indicators of acidification in aquatic ecosystems can be
chemical or biological components of the ecosystem that are altered by
the acidifying effects of nitrogen and sulfur deposition. A desirable
ecological indicator for aquatic acidification is one that is measurable
or estimable, linked causally to deposition of nitrogen and sulfur, and
linked causally, either directly or indirectly to ecological effects
known or anticipated to adversely affect public welfare.

	As summarized in chapter 2 of the Policy Assessment, atmospheric
deposition of NOy and SOx causes aquatic acidification through the input
of strong acid anions (e.g., NO3- and 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 acid neutralizing capacity (ANC); alkalinity (ALK);
base neutralizing capacity, commonly referred to as acidity (ACY);
surface water pH; concentrations of trivalent aluminum, Al+3; and
concentrations of major anions (SO42-, NO3-), cations (Ca2+, Mg+2, K+),
or sums of cations or anions.

	ANC and ALK are very similar quantities and are used interchangeably in
the literature and for some of the analyses presented in this document.
 Both ANC and ALK are defined as the amount of strong acid required to
reach a specified equivalence point.  For acid-base solutions, an
equivalence point can be thought of as the point to 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.   ANC is preferred
over ALK as the body of scientific evidence has focused on ANC and
effects relationships.   ALK is more widely associated with more general
characterizations of water quality such as the relative hardness of
water associated with carbonates.

	Indictors such as the concentrations of specific anions, cations, or
their groupings, while relevant to acidification processes, are not
robust acidification indicators as it is the relative balance of cations
and anions that is more directly associated with acidification.  That
balance is captured by ANC and ALK.   Acidity, ACY, is the corollary
coverse of ANC from the perspective of 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, aluminum, or metals in
general, have high variability in concentrations that can be linked to
effects, often at extremely low levels which in some cases approach
detectability limits, exhibit rapid transient responses, and are often
confounded by the presence of other toxic metals. These concerns limit
the use of metals as reliable and measurable ecological indicators.
Hydrogen ion (H+) concentrations, using their negative logarithmic
values, or pH, are well correlated with adverse effects, as discussed
above in section II.A, and determine the solubility of metals such as
aluminum.  However, pH is not a preferred acidification indicator due to
its highly transient nature and other concerns, as discussed below.

	Having reasoned that ANC is a preferred indicator to ALK, ACY,
individual metals or groupings of ions, the Policy Assessment considers
the relative merits of ANC compared to pH, which is a well recognized
indicator of acidity and a more direct causative agent with regard to
adverse effects.  First, the linkage between ANC and pH is considered in
recognition of the causative association between pH and effects.

	ANC is not the direct causative toxic agent impacting aquatic species
diversity.   The scientific literature generally emphasizes the links
between pH and adverse effects as described above in section II.A.   It
is important, therefore, to consider the extent to which ANC and pH are
well related from a mechanistic perspective as well as through empirical
evidence.   ANC and pH are co-dependent on each other based on the
requirement that all solutions are electrically neutral, meaning that
any solution must satisfy the condition that all negatively charged
species must be balanced by all positively charged species.   ANC is
defined as the difference between strong anions and cations (US EPA,
2011, equation 7-13).

	While the chemistry can be complex, the co-dependency between ANC and
pH is explained by recognizing that positively charged hydrogen, H+, is
incorporated in the charge balance relationships related to the overall
solution chemistry which also defines ANC.  The positive, directional
co-dependency (i.e., ANC and pH increase together) is further explained
in concept as ANC reflects how much strong acid (i.e., how much hydrogen
ion) it takes to titrate to an equivalence point.  Strong observed
correlations between pH and ANC as described in the Policy Assessment
support these mechanistic relationships.

	As discussed above in section II.A, there are well established examples
of ANC correlating strongly with a variety of ecological effects which
are summarized in the Policy Assessment (US EPA, 2011, Table 3-1).
Because pH and ANC are well correlated and linearly dependent over the
pH ranges (4.5-6) where adverse ecological effects are observed,
evidence of clear associations exist between ANC and adverse ecological
effects as described in the Policy Assessment.  In large measure, this
dependence between pH and ANC and the relationship of both pH and ANC to
effects, speak directly to the appropriateness of ANC with respect to
its use as an ecological indicator.

	Thus, there is a clear association between ANC and ecological effects,
although there is a more direct causal relationship between pH and
effects.  Nonetheless, ANC is preferred as an ecological indicator based
on its superior ability to provide a linkage with deposition in a
meaningful and quantifiable manner, a role that is served far more
effectively by ANC than by pH.  While both ANC and pH are clearly
associated with the effects of concern, ANC is superior in linking these
effects to deposition.

	The Policy Assessment notes that the basis for this conclusion is that
acidifying atmospheric deposition of nitrogen and sulfur is a direct
input of potential acidity (ACY), or, in terms of ANC, such deposition
is relevant to the major anions that reduce the capacity of a water body
to neutralize acidity.  Consequently, there is a well defined linear
relationship between potential acidifying deposition and ANC.  This
ANC-deposition relationship facilitates the linkage between ecosystem
models that calculate an ecological indicator and the atmospheric
deposition of NOy and SOx.  On the other hand, there is no direct linear
relationship between deposition and pH.  While acid inputs from
deposition lower pH, the relationship can be extremely nonlinear and
there is no direct connection from a modeling or mass balance
perspective between the amount of deposition entering a system and pH.
The term “mass balance” underlies the basic formulation of any
physical modeling construct, 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.  ANC is a conserved property.  This means
that ANC in a water body can be accounted for by knowledge of how much
ANC initially exists, how much flows in and is deposited, and how much
flows out.  In contrast, hydrogen ion concentration in the water, the
basis for pH, is not a conserved property as its concentration is
affected by several factors such as temperature, atmospheric pressure,
mixing conditions of a water body, and the levels of other several 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 Policy Assessment provides further details on why pH is not a
conserved quantity and is subject to rapid transient response behavior
that makes it difficult to use as a reliable and functional ecological
indicator.  The observed pH-to-ANC relationship (US EPA, 2011, figure
7-2) partially explains the concern with pH responding too abruptly.
In the region where pH ranges roughly from 4.5 to 6 and is of greatest
relevance to effects (US EPA, 2011, figure 7-4), there clearly is more
sensitivity of pH to changes in ANC in the ANC range from approximately
0 to 50 µeq/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 µeq/L, which is subject to very rapid change in
ANC, or deposition inputs.   This part of the relationship coincides
with reduced levels of ANC and hence with reduced ability to neutralize
acids and moderate pH fluctuations.  This response behavior can be
extended to considering how pH would change in response to deposition,
or ambient concentrations, of NOy and SOx, which can be viewed as
“ANC-like” inputs.

	In summary, because ANC clearly links both to biological effects of
aquatic acidification as well as to acidifying inputs of NOy and SOx
deposition, the Policy Assessment concludes that ANC is an appropriate
ecological indicator for relating adverse aquatic ecosystem effects to
acidifying atmospheric deposition of SOx and NOy, and is preferred to
other potential indicators.  In reaching this conclusion, the Policy
Assessment notes that in its review of the first draft Policy
Assessment, CASAC concluded that “information on levels of ANC
protective to fish and other aquatic biota has been well developed and
presents probably the lowest level of uncertainty in the entire
methodology” (Russell and Samet, 2010a).  In its more recent review of
the second draft Policy Assessment, CASAC agreed “that acid
neutralizing capacity is an appropriate ecological measure for
reflecting the effects of aquatic acidification” (Russell and Samet,
2010b; p. 4).

2.	Linking ANC to Deposition

	There is evidence to support a quantified relationship between
deposition of nitrogen and sulfur and ANC.  This relationship was
analyzed in the REA for two case study areas, the Adirondack and
Shenandoah Mountains, based on time-series modeling and observed trends.

In the REA analysis, long-term trends in surface water nitrate, sulfate
and ANC were modeled using Model of Acidification of Groundwater in
Catchment (MAGIC) for the two case study areas.  These data were used to
compare recent surface water conditions in 2006 with preindustrial
conditions (i.e. preacidification 1860).  The results showed a marked
increase in the number of acid impacted lakes, characterized as a
decrease in ANC levels, since the onset of anthropogenic nitrogen and
sulfur deposition, as discussed in chapter 2 of the Policy Assessment.

In the REA, more recent trends in ANC, over the period from 1990 to
2006, were assessed using monitoring data collected at the two case
study areas.  In both case study areas, nitrate and sulfate deposition
decreased over this time period.  In the Adirondack Mountains, this
corresponded to a decreased concentration of nitrate and sulfate in the
surface waters and an increase in ANC (US EPA, 2009, section <IP_ADDRESS>).
In the Shenandoah Mountains, there was a slight decrease in nitrate and
sulfate concentration in surface waters corresponding to modest increase
in ANC from 50 ueq/L in 1990 to 67 ueq/L in 2006 (US EPA, 2009, section
<IP_ADDRESS>,Appendix 4, and section 3.4).

	In the REA, the quantified relationship between deposition and ANC was
investigated using ecosystem acidification models, also referred to as
acid balance models or critical loads models (US EPA, 2011, section 2
and US EPA, 2009, section 4 and Appendix 4).  These models quantify the
relationship between deposition of nitrogen and sulfur and the resulting
ANC in surface waters based on an ecosystem’s inherent generation of
ANC and ability to neutralize nitrogen deposition through biological and
physical processes.  A critical load is defined as the amount of
acidifying atmospheric deposition of nitrogen and sulfur beyond which a
target ANC is not reached.   Relatively high critical load values imply
that an ecosystem can accommodate greater deposition levels than lower
critical loads for a specific target ANC level.   Ecosystem models that
calculate critical loads form the basis for linking deposition to ANC.

	As discussed in chapter 2 of the Policy Assessment, both dynamic and
steady state models calculate ANC as a function of ecosystem attributes
and atmospheric nitrogen and sulfur deposition, and can be used to
calculate critical loads.   Steady state models are time invariant and
reflect the long term consequences associated with an ecosystem reaching
equilibrium under a constant level of atmospheric deposition.   Dynamic
models are time variant and take into account the time dependencies
inherent in ecosystem hydrology, soil and biological processes.
Dynamic models like MAGIC can provide the time series response of ANC to
deposition whereas steady state models provide a single ANC relationship
to any fixed deposition level.   Dynamic models naturally are more
complex than steady state models as they attempt to capture as much of
the fundamental biogeochemical processes as practicable, whereas steady
state models depend on far greater parameterization and generalization
of processes that is afforded, somewhat, by not having to accounting for
temporal variability.

	 The Policy Assessment notes that steady state models are capable of
addressing the question of what does it take to reach and sustain a
specific level of ANC.  Dynamic models are also capable of addressing
that question, but can also address the question of how long it takes to
achieve that result.  Dynamic models afford the ability for more
comprehensive treatment of a variety of processes throughout the
surface, soil and bedrock layers within an ecosystem.  For example,
steady state models treat sulfate as a mobile anion throughout the
system, meaning that the sulfate that is deposited to a watershed enters
the water column and is not influenced by soil adsorption or cation
exchange.  Dynamic models can incorporate these time variant processes.
The use of a steady state model treating sulfate as totally mobile does
not necessarily conflict with the possibility of sulfate acting as a
less than mobile ion at certain times.   The steady state assumption is
premised on the long term behavior of sulfate which can undergo periods
of net adsorption followed by periods of net desorption which can
balance out over time.  The Policy Assessment recognizes that as the
richness of the available data increases, in terms of parameters and
spatial resolution, the incorporation of dynamic modeling approaches in
the standard setting process should become more feasible.  In
determining an appropriate modeling approach for the development of a
NAAQS in this review, the Policy Assessment considers both the relevance
of the question addressed as well as the ability to perform modeling
that provides relevant information for geographic areas across the
country.

	Dynamic models require a large amount of catchment level-specific data
relative to steady state models.  Because of the time invariant nature
of steady state models, the data requirements that integrate across a
broad spectrum of ecosystem processes is achievable and available now at
the national level.  Water quality data to support steady state models
currently exist for developing a national data base for modeling nearly
10,000 catchments in the contiguous U.S.  In contrast, the data needs to
support dynamic models for national-scale analyses simply are not
available at this time.  Further, the information provided by steady
state modeling would be sufficient to develop and analyze alternative
NAAQS and the kind of protection they would afford.  While it would be
of interest to also obtain information about how much time it would take
for a target ANC level to be achieved, the absence of such information
does not preclude developing and evaluating alternative NAAQS using the
AAI structure.  Based on the above considerations, the Policy Assessment
concludes that at this time steady state critical load modeling is an
appropriate tool for linking long-term ANC levels to atmospheric
deposition of nitrogen and sulfur for development of an AAI that has
national applicability.

	 A steady state model is used to define the critical load, which is the
amount of atmospheric deposition of nitrogen (N) and sulfur (S) beyond
which a target ANC is not achieved and sustained.  It is expressed as:

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

Where:

 CLANClim(N + S) is the critical load of deposition, with units of
equivalent charge/(area-time);

 [BC]0* is the natural contribution of base cations from weathering,
soil processes and preindustrial deposition, with units of equivalent
charge/volume;

[ANClim] is the target ANC value, with units of equivalent
charge/volume;

Q is the catchment level runoff rate governed by water mass balance and
dominated by precipitation, with units of distance/time; and

Neco is the amount of nitrogen deposition that is effectively
neutralized by a variety of biological (e.g., nutrient uptake) and
physical processes, with units of equivalent charge/ (area-time).

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

	In considering the contributions of SOx or NOy species to
acidification, it is useful to think of every depositing nitrogen atom
as supplying one equivalent charge unit and every sulfur atom as
depositing two charge units.  The Policy Assessment uses equivalent
charge per volume as a normalizing tool in place of the more familiar
metrics such as mass or moles per volume.  This allows for a clearer
explanation of many of the relationships between atmospheric and
ecosystem processes that incorporate mass and volume unit conventions
somewhat specific to the environmental media of concern (e.g., m3 for
air and liter for liquid water).  Equivalent charge reflects the
chemistry equilibrium fundamentals that assume electroneutrality, or
balancing charge where the sum of cations always equals the sum of
anions.

	As presented above, the terms S and N in the CLANClim (N + S) term
broadly represent all species of sulfur or nitrogen that can contribute
to acidifying deposition.  This follows conventions used in the
scientific literature that addresses critical loads, and it reflects all
possible acidifying contributions from any sulfur or nitrogen species.
For all practical purposes, S reflects SOx as described above, the sum
of sulfur dioxide gas and particulate sulfate.   However, N in equation
III-1 includes both oxidized forms, consistent with the ambient
indicator, NOy, in addition to the reduced nitrogen species, ammonia and
ammonium ion, referred to as NHx.   NHx is included in the critical load
formulation because it contributes to potentially acidifying nitrogen
deposition.  Consequently, from a mass balance or modeling perspective,
the form of the standard needs to account for NHx, as described below.

3.	Linking Deposition to Ambient Air Indicators

	The last major component of the form illustrated in Figure III-1
addresses the linkage between deposition of nitrogen and sulfur and
concentrations of the ambient air indicators, NOy and SOx.   To link
ambient air concentrations with deposition, the Policy Assessment
defines a transference ratio, T, as the ratio of total wet and dry
deposition to ambient concentration, consistent with the area and time
period over which the standard is defined.  To express deposition of NOy
and SOx in terms of NOy and SOx ambient concentrations, two transference
ratios were defined, where TSOx equals the ratio of the combined dry and
wet deposition of SOx to the ambient air concentration of SOx, and TNOy
equals the ratio of the combined dry and wet deposition of NOy to the
ambient air concentration of NOy.

	As described in Chapter 7 of the Policy Assessment, reduced forms of
nitrogen (NHX) are included in total nitrogen in the critical load
equation, III-1.  Reduced forms of nitrogen are treated separately, as
are NOy and SOx, and the transference ratios are applied.  This results
in the following critical load expression that is defined explicitly in
terms of the indicators NOy and SOx:

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

This is the same equation as III-1, with the deposition associated with
the critical load translated to deposition from ambient air
concentrations via transference ratios.  In addition, deposition of
reduced nitrogen, oxidize nitrogen and oxidized sulfur are treated
separately.

	Transference ratios are a modeled construct, and therefore cannot be
compared directly to measurable quantities.  There is an analogy to
deposition velocity, as a transference ratio is basically an aggregated
weighted average of the deposition velocities of all contributing
species across dry and wet deposition, and transference ratio units are
expressed as distance/time.  However, wet deposition commonly is not
interpreted as the product of a concentration times a velocity.  Direct
wet deposition observations are available which integrate all of the
processes, regardless of how well they may be understood, related to wet
deposition into a measurable quantity.  There are reasonable analogies
between the processes governing dry and wet deposition, from a
fundamental mass transfer perspective.  In both cases there is a
transfer of mass between the dry ambient phase and another medium,
either a surface or vegetation in the case of dry deposition, or a rain
droplet or cloud in the case of wet precipitation. The specific
thermodynamic properties and chemical/biological reactions that govern
the transfer of dry mass to plants or aqueous droplets differ, but
either process can be based on conceptualizing the product of a
concentration, or concentration difference, times a mass transfer
coefficient which is analogous to the basic dry deposition model:  dry
deposition = concentration x velocity (US EPA, 2011, Appendix F).

	Transference ratios require estimates of wet deposition of NOy and SOx,
dry deposition of NOy and SOx, and ambient air concentrations of NOy and
SOx.  Possible sources of information include model estimates or a
combination of model estimates and observations, recognizing that dry
deposition is a modeled quantity that can use observed or modeled
estimates of concentration.  The limited amount of NOy measurements in
acid-sensitive areas as well as the combination of representative NOy,
SO2 and SO4 observations generally preclude the use of observations for
development of a standard that is applicable nationally.

	The Policy Assessment considers a blending of observations and models
to take advantage of their relative strengths; e.g., combining the NADP
wet deposition observations, modeled dry deposition, and a mix of
modeled and observed concentrations, using the model for those species
not measured or measured with very sparse spatial coverage.  A potential
disadvantage of mixing and matching observations and model estimates is
to lose consistency afforded by using just modeling alone.  A modeling
platform like CMAQ is based on adhering to consistent treatment of mass
conservation, by linking emission inputs with air concentrations and
concentrations to deposition.  Inconsistencies from combining processes
from different analytical platforms increase the chance that mass (of
nitrogen or sulfur) would unintentionally be increased or decreased as
the internal checking that assures mass conservation is lost.
Transference ratios incorporate a broad suite of atmospheric processes
and consequently an analytical approach that instills consistency in the
linkage of these processes is preferable to an approach lacking such
inherent consistency.  This contention does not mean that observations
alone, if available, could not be used, but suggests that the
inconsistencies in combining models and observations for the purposes of
developing transference ratios has the potential for creating unintended
artifacts.

	While there is a reasonable conceptual basis for the concept of an
aggregated deposition velocity referred to in the Policy Assessment as a
transference ratio, there is very limited ability to compare observed
and calculated ratios.  This is because the deposition velocity is
dependent on individual species, and the mass transfer processes of wet
and dry removal, while conceptually similar, are different.
Consequently, there does not exist a meaningful approach to measure such
an aggregated or lumped parameter.  Therefore, at this time, the
evaluation of transference ratios is based on sensitivity studies,
analysis of variability, and comparisons with other models, as described
in Appendix F of the Policy Assessment.

	As discussed in Appendix F, the interannual variability, as well as the
sensitivity to emission changes of roughly 50%, results in changes of
transference ratios of approximately 5 - 10%.  Part of the reason for
this inherent stability is due to the co-dependence of concentration and
deposition.  For example, as concentrations are reduced as a result of
emissions reductions, deposition in turn is reduced since deposition is
a direct linear function of concentration leading to negligible impact
on the deposition-to-concentration ratio.  Likewise, an overestimate of
concentration likely does not induce a bias in the transference ratio.
While it is important to continue to improve the model’s ability to
match ambient concentrations in time and space, the bias of a modeled
estimate of concentration relative to observations does not necessarily
result in a bias in a calculated transference ratio.  In effect, this
consideration of bias cancellation reduces the sensitivity of
transference ratios to model uncertainties and affords increased
confidence in the stability of these ratios.  Based on the series of
sensitivity and variability analyses, the Policy Assessment concludes
that the transference ratios are relatively stable and provide a sound
metric for linking deposition and concentration in the form of the
standard.

in species definitions and spatial configurations, it does suggest two
very important conclusions.  First, the idea of using multiple platforms
for different parts of the country may be problematic as there does not
exist a reliable approach to judge acceptance which is almost always
based on comparisons to observations.  Second, since transference ratios
are based on concentrations and deposition, as the uncertainties in each
of those components are reduced, the relative uncertainty in the ratios
also is reduced.  This means that basic improvements in the model’s
ability to reproduce observed wet deposition and ambient concentration
fields enhance the relative confidence in the constructed transference
ratios.  Similarly, as in-situ dry deposition flux measurements become
available that enable a more rigorous evaluation and diagnosis of
modeled dry deposition processes, the expected improved treatment of dry
deposition also would increase confidence in transference ratios.
Finally, deposition is directly related to ambient air concentrations.
Models like CMAQ rely on the concentration-to-deposition linkage to
calculate deposition, which is the foundation for broadly based and
robust assessments addressing atmospheric deposition.  In principle, the
use of a modeled constructed transference ratio is based on the same
premise by which we use models to estimate deposition in the first
place.

	The shortage of widely available ambient air observations and the fact
that estimates of dry deposition requires modeling, collectively
suggests that a unified modeling platform is the best approach for
constructing transference ratios.  The Policy Assessment (US EPA,
2011,section 2) considers CMAQ and other models, such as CAMx and the
Canada’s AURAMS - A Unified Regional Air-quality Modeling System
(Smythe et al., 2008), and concludes that CMAQ is the preferred modeling
platform for constructing transference ratios for purposes of developing
a new secondary standard for consideration.  This conclusion reflects
the view that for the purposes of defining transference ratios, a
modeling platform should (1) be a multiple pollutant model recognizing
the myriad of connections across pollutant categories that directly and
indirectly impact nitrogen and sulfur characterization, (2) include the
most comprehensive scientific treatments of atmospheric processes that
relate directly and indirectly to characterizing concentrations and
deposition, (3) have an infrastructure capability that accommodates the
inclusion of improved scientific treatments of relevant processes and
important input fields, and (4) undergo frequent reviews regarding the
adequacy of the underlying science as well as the appropriateness in
applications.  The CMAQ platform exhibits all these characteristics.  It
has been (and continues to be) extensively evaluated for several
pollutant categories, is supported by a central infrastructure of EPA
scientists, with considerable interfacing with the scientific research
communities in academia and industry,  whose mission is to improve and
evaluate the CMAQ platform.  More directly, CMAQ, and its predecessor
versions, has a long track record going back to the NAPAP in the
1980’s of specific improvements in deposition processes, which are
described in Appendix F of the Policy Assessment.

4.	Aquatic Acidification Index

	Having established the various expressions that link atmospheric
deposition of nitrogen and sulfur to ANC and the transference ratios
that translate atmospheric concentrations to deposition of nitrogen and
sulfur, the Policy Assessment derives the following expression of these
linkages, which separates reduced forms of nitrogen, NHx, from oxidized
forms:

 - TNOy [NOy]/Qr - TSOx[SOx]/Qr 		(III-3)

Based on equation III- 3, the Policy Assessment defines an aquatic
acidification index (AAI) that is more simply stated using terms that
highlight the ambient air indicators:

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

where the AAI represents the long term (or steady state) ANC level
associated with ambient air concentrations of NOy and SOx.   The factors
F1 through F4 convey three attributes:  a relative measure of the
ecosystem’s ability to neutralize acids (F1), the acidifying potential
of reduced nitrogen deposition (F2), and the deposition-to-concentration
translators for NOy (F3) and SOx (F4).

Specifically:

F1 = ANClim + CLr/Qr ;

F2 =  NHx/Qr  = NHx deposition divided by Qr;

F3 =  TNOy/ Qr ; TNOy is the transference ratio that converts ambient
air concentrations of NOy to deposition of NOy; and

F4 =  TSOx/ Qr ; TSOx is the transference ratio that converts ambient
air concentrations of SOx to deposition of SOx.

All of these factors include representative Qr to maintain unit (and
mass) consistency between the AAI and the terms on the right side of
equation III-4.

	The F1 factor is the target ANC level plus the amount of deposition
(critical load) the ecosystem can receive and still achieve the target
level.  It incorporates an ecosystem’s ability to generate acid
neutralizing capacity through base cation supply ([BC]*0) and to
neutralize acidifying nitrogen deposition through Neco, both of which
are incorporated in the CL term.  As noted above, because Neco can only
neutralize nitrogen deposition (oxidized or reduced) there may be rare
cases where Neco exceeds the combination of reduced and oxidized
nitrogen deposition.  Consequently, to ensure that the AAI equation is
applicable in all cases that may occur, equation III-4 is conditional on
total nitrogen deposition, {NHx + F3[NOy]}, being greater than Neco.  In
rare cases where Neco is greater than {NHx + F3[NOy]}, F2, are F3, and
Neco would be set equal to 0 in the AAI equation.  The consequence of
setting F2 and F3 to zero is simply to constrain the AAI calculation
just to SOx, as nitrogen would have no bearing on acidifying
contributions in this case.

	The Policy Assessment concludes that equation III- 4 (US EPA,
2011,equation 7-12), which defines the AAI, is ecologically relevant and
appropriate for use as the form of a national standard designed to
provide protection for aquatic ecosystems from the effects association
with acidifying deposition associated with concentrations of oxides of
nitrogen and sulfur in the ambient air.  This AAI equation does not,
however, in itself, define the spatial areas over which the terms of the
equation would apply.  To specify values for factors F1 through F4, it
is necessary to define spatial areas over which these factors are
determined.  Thus, it is necessary to identify an approach for spatially
aggregating water bodies into ecologically meaningful regions across the
U.S., as discussed below.

5.	Spatial Aggregation

	As discussed in the Policy Assessment, one of the unique aspects of
this form is the need to consider the spatial areas over which values
for the F factors in the AAI equation are quantified.  Ecosystems across
the U.S. exhibit a wide range of geological, hydrological and vegetation
characteristics that influence greatly the ecosystem parameters, Q, BC0*
and Neco that are incorporated in the AAI.  Variations in ecosystem
attributes naturally lead to wide variability in the sensitivities of
water bodies in the U.S. to acidification, as well as in the
responsiveness of water bodies to changes in acidifying deposition.
Consequently, variations in ecosystem sensitivity, and the uncertainties
inherent in characterizing these variations, must be taken into account
in developing a national standard.  In developing a secondary NAAQS to
protect public welfare, the focus of the Policy Assessment is on
protecting sensitive populations of water bodies, not on each individual
water body, which is consistent with the Agency’s approach to
protecting public health through primary NAAQS that focus on susceptible
populations, not on each individual.

	The approach used for defining ecologically relevant regions across the
U.S. in the Policy Assessment (US EPA, 2011,section 7.2.5) is described
below, along with approaches to characterizing each region as acid
sensitive or relatively non-acid sensitive.  This characterization
facilitates a more detailed analysis and focus on those regions that are
relatively more acid sensitive.  This characterization is also used to
avoid over-protection in relatively non-acid sensitive regions, regions
that would receive limited benefit from reductions in the deposition of
oxides of nitrogen and sulfur with respect to aquatic acidification
effects.  Approaches to developing representative values for each of the
terms in the AAI equation (factors F1 through F4) for each ecologically
relevant region for which sufficient data are available are then
discussed.  These spatial aggregation approaches are generally
applicable to the contiguous United States.  The following discussion
also addresses the development of factors for data-limited regions and
specifically for Hawaii, Alaska and the U.S. territories.

	Stated more simply, this section discusses appropriate ways to divide
the country into ecologically relevant regions; to characterize each
region as either acid sensitive or relatively non-acid sensitive; and to
determine values of factors F1 through F4 for each region, taking into
consideration the acid sensitivity of each region and the availability
of relevant data.  For each such region, the AAI would be calculated
based on the values of factors F1 through F4 specified for that region.

	In considering approaches to spatial aggregation, the Policy Assessment
focuses on methods that have been developed to define ecologically
relevant regions, referred to as ecoregions, which are meaningfully
related to the factors that are relevant to aquatic acidification.  As
noted above, the Policy Assessment did not focus on looking at each
individual water body, nor did it focus on aggregating over the entire
nation, which would preclude taking into account the inherent
variability in atmospheric and ecological factors that fundamentally
modify the relationships that are central to the development of an
ecologically relevant AAI.

	Based on considering available classification schemes, the Policy
Assessment concludes that Omernik’s ecoregion classification is the
most appropriate method to consider for the purposes of this review.
This classification offers several levels of spatial delineation, has
undergone an extensive scientific peer review process, and has
explicitly been applied to delineating acid sensitive areas within the
U.S.  Further, the Policy Assessment concludes that ecoregion level III
(Figure III-1) resolution, with 84 defined ecoregions in the contiguous
U.S., is the most appropriate level to consider for this purpose.  The
spatial resolution afforded by level III strikes an appropriate balance
relative to the reasoning that supports conclusions on indicators, as
discussed above.  The Policy Assessment concludes that the most detailed
level of resolution (level IV) is not appropriate given the limited data
availability to address nearly 1000 subdivisons within that level and
the currently evolving nature of level IV regions.  Further, level III
ecoregions are preferred to level II in that level III ecoregions, but
not level II ecoregions, are largely contiguous in space which allows
for a more coherent development of information to quantify the AAI
factors and to characterize the concentrations of NOy and SOx in the
ambient air within each ecoregion.

	Appendix C of the Policy Assessment includes a description of each
level III ecoregion.  The Policy Assessment notes that the use of
ecoregions is an appropriate spatial aggregation scheme for this
standard focused on deposition-related aquatic acidification effects,
while many of the same ecoregion attributes may be applicable in
subsequent NAAQS reviews that may address other deposition-related
aquatic and terrestrial ecological effects.   Because atmospheric
deposition is modified by ecosystem attributes, the types of vegetation,
soils, bedrock geology, and topographic features that are the basis of
this ecoregion classification approach also will likely be key
attributes for other deposition-related effects (e.g., terrestrial
acidification, nutrient enrichment) that link atmospheric concentrations
to an aquatic or terrestrial ecological indicator.

Figure III-1.  Omernik Ecoregion III areas ( HYPERLINK
"http://www.epa.gov/wed/pages/ecoregions"
http://www.epa.gov/wed/pages/ecoregions ).

a.	Ecoregion Sensitivity

	The Policy Assessment used Omernik’s original alkalinity data (US
EPA, 2011,as section  2) and more recent ANC data to delineate two broad
groupings of ecoregions:  acid-sensitive and relatively non-acid
sensitive ecoregions.  This delineation was made to facilitate greater
focus on those ecoregions with water bodies that generally have greater
acid sensitivity and to avoid over-protection in regions with generally
less sensitive water bodies.  The approach used to delineate
acid-sensitive and relatively non-acid sensitive regions included an
initial numerical-based sorting scheme using ANC data, which categorized
ecoregions with relatively high ANC values as being relatively non-acid
sensitive.  This initial delineation resulted in 29 of the 84 Omernik
ecoregions being categorized as acid sensitive.  Subsequently, land use
data were also considered to determine to what extent an ecoregion is of
a relatively pristine and rural nature by quantifying the degree to
which active management practices related to development and agriculture
occur in each ecoregion.

 In step 2, land use data were used to identify those acid sensitive
ecoregions with significant managed areas that would not be considered
as having a  relatively pristine and rural character.  The percentage of
the combination of developed (residential, transportation, industrial
and commercial) and agricultural (croplands, pastures, orchards,
vineyards) land use was used as an indicator of managed land use area.
Forest cover was used as an indicator of non-managed land use more
directly reflecting the pristine quality of a region.  Based on the 2006
National Land Cover Data base (NLCD,  HYPERLINK
"http://www.epa.gov/mrlc/nlcd-2006.html"
http://www.epa.gov/mrlc/nlcd-2006.html ), acid sensitive ecoregions
would meet both of the following land use data criteria: percent of
developed and agricultural area less than 20% combined with forested
area greater than 50%.  The combination of steps 1 and 2 identify 22
relatively acid sensitive areas (Table III-1 and Figure III-2).

	Consideration was also given to the use of naturally acidic conditions
in defining relatively non-acid sensitive areas.  For example, several
of the ecoregions located in plains near the coast exhibit elevated
dissolved organic carbon (DOC) levels, which is associated with
naturally acidic conditions.  DOC in surface waters is derived from a
variety of weak organic acid compounds generated from the natural
availability and decomposition of organic matter from biota.
Consequently, high DOC is associated with “natural” acidity, with
the implication that a standard intended to protect against atmospheric
contributions to acidity is not an area of focus.  The evidence suggests
that several of the more highly managed ecoregions in coastal or near
coastal transition zones are associated with relatively high DOC values,
typically exceeding on average 5 mg/l, compared to other acid sensitive
areas.  Although there is sound logic to interpret naturally acidic
areas as relatively non-acid sensitive, natural acidity indicators were
not explicitly included in defining relatively non-acid sensitive areas
as there does not exist a consensus-based quantifiable scientific
definition of natural acidity.  Approaches to explicitly define natural
acidity likely will be pursued in future reviews of the standard.

Figure III-2.  Acid-sensitive ecoregions identified in grey fill (22 out
of 84 ecoregions).

b.	Representative Ecoregion-specific Factors

	Having concluded that the Omernik level III ecoregions are an
appropriate approach to spatial aggregation for the purpose of a
standard to address deposition-related aquatic acidification effects,
the Policy Assessment uses those ecoregions to define each of the
factors in the AAI equation.  As discussed below, factors F1 through F4
in equation III-4 are defined for each ecoregion by specifying
ecoregion-specific values for each factor based on monitored or modeled
data that are representative of each ecoregion.

i.	Factor F1

As discussed above, factor F1 reflects a relative measure of an
ecosystem’s ability to neutralize acidifying deposition, and is
defined as:  F1 = ANClim + CLr/Qr.  The value of F1 for each ecoregion
would be based on a representative critical load for the ecoregion (CLr)
associated with a single national target ANC level (ANClim, discussed
below in section III.D), as well as on a representative runoff rate
(Qr).  To specify ecoregion-specific values for the term Qr, the Policy
Assessment used the median value of the distribution of Q values that
are available for water bodies within each ecoregion.  To specify
ecoregion-specific representative values for the term CLr in factor F1,
a distribution of calculated critical loads was created for the water
bodies in each ecoregion for which sufficient water quality and
hydrology data are available.  The representative critical load was then
defined to be a specific percentile of the distribution of critical
loads in the ecoregion. Thus, for example, using the 90th percentile
means that within an ecoregion, 90 percent of the water bodies would be
expected to have higher calculated critical loads than the
representative critical load.  That is, if the representative critical
load were to occur across the ecoregion, 90 percent of the water bodies
would be expected to achieve the national ANC target or better.

The specific percentile selected as part of the definition of F1 is an
important parameter that directly impacts the representative critical
load specified for each ecoregion, and therefore the degree of
protectiveness of the standard.   A higher percentile corresponds to a
lower critical load and, therefore, to lower allowable ambient air
concentrations of NOy and SOx and related deposition to achieve a target
AAI level.  In conjunction with the other terms in the AAI equation,
alternative forms can be appropriately characterized in part by
identifying a range of alternative percentiles.  The choice of an
appropriate range of percentiles to consider for acid-sensitive and
relatively non-acid sensitive ecoregions, respectively, is discussed
below.

(a)	Acid-sensitive Ecoregions

critical load, so as to avoid potential extreme outliers that can be
seen to exist at the extreme end of the data distributions, which would
not be representative of the population of acid sensitive water bodies
within the ecoregion and could lead to an overly protective standard.
Also At the same time, in considering ecoregions that are inherently
acid sensitive, it is judged to be appropriate to limit the lower end of
the range for consideration to the 70th percentile, a value well above
the median of the distribution, so that a substantial majority of
acid-sensitive water bodies are protected.

	In considering this conclusion, the CASAC Panel noted that the data
bases for calculating critical loads within an ecoregion are not
necessarily representative of all water bodies within an ecoregion.
That is, in many ecoregions the lake sampling design used in studies
that generated the relevant data may have focused on the relatively more
sensitive water bodies within an ecoregion (Russell and Samet, 2011a).
Consequently, a given percentile of the distribution of calculated
critical loads, based on sampled water bodies, may not be representative
of that percentile of all water bodies across an entire ecoregion.  To
the extent that the sampling of water bodies within an ecoregion was
skewed toward the relatively more sensitive water bodies, selecting a
given percentile from the distribution of available critical loads would
result in a somewhat higher percentile of all water bodies within that
ecoregion having a higher calculated critical load than the
representative critical load value.  Thus, the extent to which study
sampling designs have resulted in skewed distributions of calculated
critical loads is an uncertainty that is appropriate to consider in
selecting a percentile for the purpose of defining the factor F1 in the
AAI equation.

(b)	Non-acid sensitive Ecoregions

	With regard to identifying percentiles that are appropriate to consider
for the purpose of calculating factor F1 for ecoregions characterized as
relatively non-acid sensitive, the Policy Assessment recognizes that
while such ecoregions are generally less sensitive to acidifying
deposition from oxides of nitrogen and sulfur, they may contain a number
of water bodies that are acid sensitive   This category includes
ecoregions that are well protected from acidification effects due to
natural production of base cations and high ANC levels, as well as
naturally acidic systems with limited base cation production and
consequently very low critical loads.  Therefore, the use of a critical
load that would be associated with a highly sensitive water body in a
naturally acidic system would impose a high degree of relative
protection in terms of allowable ambient air concentrations of oxides of
nitrogen and sulfur and related deposition, while potentially affording
little or no public welfare benefit from attempting to improve a
naturally acidic system.

ii.	Factor F2

	As discussed above, factor F2 is the amount of reduced nitrogen
deposition within an ecoregion, including the deposition of both ammonia
gas and ammonium ion, and is defined as:  F2 = NHx/Qr.  The Policy
Assessment calculated the representative runoff rate, Qr, using a
similar approach as noted above for factor F1; i.e., the median value of
the distribution of Q values that are available for water bodies within
each ecoregion. In the Policy Assessment, 2005 CMAQ model simulations
over 12-km grids are used to calculate an average value of NHx for each
ecoregion. The NHx term is based on annual average model outputs for
each grid cell, which are spatially averaged across all the grid cells
contained in each ecoregion to calculate a representative annual average
value for each ecoregion. The Policy Assessment concludes that this
approach of using spatially averaged values is appropriate for modeling,
largely due to the relatively rapid mixing of air masses that typically
results in relatively homogeneous air quality patterns for regionally
dispersed pollutants.  In addition, there is greater confidence in using
spatially averaged modeled atmospheric fields than in using modeled
point-specific fields.

	This averaging approach is also used for the air concentration and
deposition terms in factors F3 and F4, as discussed below.  The Policy
Assessment notes that modeled NHx deposition exhibits greater spatial
variability than the other modeled terms in factors F3 and F4.
Recognizing this greater variability, the Policy Assessment concludes
that it would be appropriate to consider alternative approaches to
specifying the value of NHx.  One such approach might involve the use of
more localized and/or contemporaneous modeling in areas where this term
is likely to be particularly variable and important.

iii.	Factors F3 and F4

	As discussed above, factors F3 and F4 are the ratios that relate
ambient air concentrations of NOy and SOx to the associated deposition,
and are defined as follow:  F3 =  TNOy/ Qr  and F4 =  TSOx/ Qr.  TNOy is
the transference ratio that converts ambient air concentrations of NOy
to deposition of NOy and TSOx is the transference ratio that converts
ambient air concentrations of SOx to deposition of SOx.  The
representative runoff rate, Qr, is calculated as for factors F1 and F2.
The transference ratios are based on the 2005 CMAQ simulations, using
average values for each ecoregion, as noted above for factor F2.  More
specifically, the transference ratios are calculated as the annual
deposition of NOy or SOx spatially averaged across the ecoregion and
divided by the annual ambient air concentration of NOy or SOx,
respectively, spatially averaged across the ecoregion.

c.	Factors in Data-limited Ecoregions

	As discussed above in section III.B.5.a, in the Policy Assessment the
initial delineation of acid-sensitive and relatively non-acid sensitive
ecoregions was based on available ANC and alkalinity data.   Areas not
meeting the ANC criteria described above are categorized as relatively
non-acid sensitive.  The development of a reasonable distribution of
critical loads for water bodies within an ecoregion for the purpose of
identifying the representative critical load requires additional data,
including more specific water quality data for major cations and anions.
 This means that the water bodies that can be used to develop a
distribution of critical loads is generally a subset of those water
bodies for which ANC data are available   Consequently, there are
certain ecoregions with sparse data that are not suitable for developing
a distribution of critical loads.

	As noted above, the Policy Assessment judges that it is not appropriate
to develop such distributions based on data from less than ten water
bodies within an ecoregion.  Twelve such ecoregions, which included only
relatively non-acid sensitive ecoregions, were characterized as being
data-limited.  For these ecoregions, the Policy Assessment considered
alternative approaches to specifying values for the terms CLr and Qr for
the purpose of determining values for each of the factors in the AAI
equation.  For these data-limited ecoregions, the Policy Assessment
judges that it is appropriate to use the median values of CLr and Qr
from the distributions of these terms for all other relatively non-acid
sensitive ecoregions, rather than attempting to use severely limited
data to develop a value for these terms based solely on data from such
an ecoregion.  Further, consideration could be given to using a single
national default value for all relatively non-acid sensitive ecoregions.
 The Policy Assessment notes that this data limitation is not a concern
in specifying values for the other terms in the AAI equation for such
ecoregions, since those terms are based on data from the 2005 CMAQ model
simulation, which covers all ecoregions across the contiguous U.S.

d.	Application to Hawaii, Alaska, and the U.S. Territories

	The above methods for specifying ecoregion-specific values for the
factors in the AAI equation apply to those ecoregions within the
contiguous U.S.  For areas outside the continental U.S., including
Hawaii, Alaska, and the U.S. Territories, there is currently a lack of
available data to characterize the sensitivity of such areas, as well as
a lack of water body-specific data and CMAQ-type modeling to specify
values for the F1 through F4 factors.   Thus, the Policy Assessment has
considered possible alternative approaches to specifying values for
factors F1 through F4 in the AAI equation for these areas.

	One such approach could be to specify area-specific values for the
factors based on values derived for ecoregions with similar acid
sensitivities, to the extent that relevant information can be obtained
to determine such similarities.  Such an approach would involve
conducting an analysis to characterize similarities in relevant
ecological attributes between ecoregions in the contiguous U.S. and
these areas outside the contiguous U.S. so as to determine the
appropriateness of utilizing ecoregion-specific values for the CLr and
Qr terms from one or more ecoregions within the contiguous U.S.  This
approach would also involve conducting additional air quality modeling
for these area that are outside the geographical scope of the currently
available CMAQ model simulations, so as to develop the other information
necessary to specify values for factors F2 through F4 for these areas.

	A second approach could rely on future data collection efforts to
establish relevant ecological data within these areas that, together
with additional air quality modeling, could be used to specify
area-specific values for factors F1 through F4.  Until such time as
relevant data become available, these areas could be treated the same as
data-limited ecoregions in the contiguous U.S. that are relatively
non-acid sensitive.

	The Policy Assessment concludes that either approach would introduce
substantial uncertainties that arise from attempting to extrapolate
values based on similarity assumptions or arbitrarily assigning values
for factors in the AAI equation that would be applicable to these areas
outside the contiguous U.S.  In light of such uncertainties, the Policy
Assessment concludes that it would also be appropriate to consider
relying on the existing NO2 and SO2 secondary standards in these areas
for protection of any potential direct or deposition-related ecological
effects that may be associated with the presence of oxides of nitrogen
and sulfur in the ambient air.  The Policy Assessment concludes that
relying on existing secondary standards in these areas is preferable to
using a highly uncertain approach to allow for the application of a new
standard based on the AAI in the absence of relevant area-specific data.

6.	Summary of the AAI Form

	With regard to the form of a multi-pollutant air quality standard to
address deposition-related aquatic acidification effects, the Policy
Assessment concludes that consideration should be given to an
ecologically relevant form that characterizes the relationships between
the ambient air indicators for oxides of nitrogen and sulfur, the
related deposition of nitrogen and sulfur, and the associated aquatic
acidification effects in terms of a relevant ecological indicator.
Based on the available information and assessments, consideration should
be given to using acid neutralizing capacity (ANC) as the most
appropriate ecological indicator for this purpose, in that it provides
the most stable metric that is highly associated with the water quality
properties that are directly responsible for the principal adverse
effects associated with aquatic acidification:  fish mortality and
reduced aquatic species diversity.

	The Policy Assessment developed such a form, termed an aquatic
acidification index (AAI), using a simple equation to calculate an AAI
value in terms of the ambient air indicators of oxides and nitrogen and
sulfur and the relevant ecological and atmospheric factors that modify
the relationships between the ambient air indicators and ANC.
Recognizing the spatial variability of such factors across the U.S., the
Policy Assessment concludes it is appropriate to divide the country into
ecologically relevant regions, characterized as acid-sensitive or
relatively non-acid-sensitive, and specify the value of each of the
factors in the AAI equation for each such region.  Omernik ecoregions,
level III, are identified as the appropriate set of regions over which
to define the AAI.  There are 84 such ecoregions that cover the
continental U.S.  This set of ecoregions is based on grouping a variety
of vegetation, geological, and hydrological attributes that are directly
relevant to aquatic acidification assessments and that allow for a
practical application of an aquatic acidification standard on a national
scale.

	The Policy Assessment defines AAI by the following equation:  AAI = F1
– F2 – F3[NOy] – F4[SOx].  Factors F1 through F4 would be defined
for each ecoregion by specifying ecoregion-specific values for each
factor based on monitored or modeled data that are representative of
each ecoregion.  The F1 factor is also defined by a target ANC value.
More specifically:

	(1)  F1 reflects a relative measure of an ecosystem’s ability to
neutralize acidifying 	deposition.  The value of F1 for each ecoregion
would be based on a representative 	critical load for the ecoregion
associated with a single national target ANC level, as well 	as on a
representative runoff rate.  The representative runoff rate, which is
also used in 	specifying values for the other factors, would be the
median value of the distributions of 	runoff rates within the ecoregion.
 The representative critical load would be derived from 	a distribution
of critical loads calculated for each water body in the ecoregion for
which 	sufficient water quality and hydrology data are available.  The
representative critical load 	would be defined by selecting a specific
percentile of the distribution.

		In identifying a range of percentiles that are appropriate to consider
for this 	purpose, regions categorized as acid sensitive were considered
separately from regions 	categorized as relatively non-acid sensitive.
For acid sensitive regions, the Policy 	Assessment concludes that
consideration should be given to selecting a percentile from 	within the
range of the 70th to the 90th percentile.  The lower end of this range
was 	selected to be appreciably above the median value so as to ensure
that the critical load 	would be representative of the population of
relatively more acid sensitive water bodies 	within the region, while
the upper end was selected to avoid the use of a critical load 	from the
extreme tail of the distribution which is subject to a high degree of
variability 	and potential outliers.  For relatively non-acid sensitive
regions, the Policy Assessment 	concludes that consideration should be
given to selecting the 50th percentile to best 	represent the
distribution of water bodies within such a region, or alternatively to
using 	the median critical load of all relatively non-acid sensitive
areas, recognizing that such 	areas are far less frequently evaluated
than acid sensitive areas.  Using either of these 	approaches would
avoid characterizing a generally non-acid-sensitive region with a
critical load that is representative of relatively acid sensitive water
bodies that may exist 	within a generally non-acid sensitive region.

(2)  F2 reflects the deposition of reduced nitrogen.  Consideration
should be given to specifying the value of F2 for each region based on
the averaged modeled value across the region, using national CMAQ
modeling that has been conducted by EPA.  Consideration could also be
given to alternative approaches to specifying this value, such as the
use of more localized and/or contemporaneous modeling in areas where
this term is likely to be particularly variable and important.

(3)  F3 and F4 reflect transference ratios that convert ambient air
concentrations of NOy and SOx, respectively, into related deposition of
nitrogen and sulfur.  Consideration should be given to specifying the
values for F3 and F4 for each region based on CMAQ modeling results
averaged across the region.  We conclude that specifying the values or
the transference rations based on CMAQ modeling results alone is
preferred to an alternative approach that combines CMAQ model estimates
with observational data.

(4)  The terms [NOy] and [SOx] reflect ambient air concentrations
measured at monitoring sites within each region.

	Using the equation, a value of AAI can be calculated for any measured
values of ambient NOy and SOx.  For such a NAAQS, the Administrator
would set a single, national value for the level of the AAI used to
determine achievement of the NAAQS, as discussed below in section III.D.
 The ecoregion-specific values for factors F1 through F4 would be
specified by EPA based on the most recent data and CMAQ model
simulations, and codified as part of such a standard.  These factors